United States
Department
of Agriculture
Forest Service
Rocky Mountain
Research Station
General Technical
Report RMRS-GTR-174
September 2006
Strategies for Conserving Native
Salmonid Populations at Risk
From Nonnative Fish Invasions:
Tradeoffs in Using Barriers to Upstream Movement
Kurt D. Fausch
Bruce E. Rieman
Michael K. Young
Jason B. Dunham
Fausch, Kurt D.; Rieman, Bruce E.; Young, Michael, K.; Dunham, Jason B. 2006. Strategies for
conserving native salmonid populations at risk from nonnative fish invasions: tradeoffs
in using barriers to upstream movement. Gen. Tech. Rep. RMRS-GTR-174. Fort Collins, CO:
U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 44 p.
Abstract_ _________________________________________
Native salmonid populations in the inland West are often restricted to small isolated habitats at
risk from invasion by nonnative salmonids. However, further isolating these populations using barriers to prevent invasions can increase their extinction risk. This monograph reviews the state of
knowledge about this tradeoff between invasion and isolation. We present a conceptual framework
to guide analysis, focusing on four main questions concerning conservation value, vulnerability
to invasion, persistence given isolation, and priorities when conserving multiple populations. Two
examples illustrate use of the framework, and a final section discusses opportunities for making
strategic decisions when faced with the invasion-isolation tradeoff.
Keywords: barriers to fish movement, biological invasions, conservation biology, isolation
­management, native salmonids
Authors___________________________________________________
Kurt D. Fausch is a Professor in the Department of Fish, Wildlife, and Conservation Biology at Colorado
State University, in Fort Collins, CO. He earned a B.S. from the University of Minnesota, and M.S. and
Ph.D. degrees from Michigan State University. During the last 30 years his research has focused on
conservation and management of salmonids and other stream fishes, throughout the West and several
locations worldwide.
Bruce E. Rieman is a Research Fisheries Biologist with the Rocky Mountain Research Station Boise
Aquatic Sciences Laboratory in Boise, ID. He holds B.S., M.S. and Ph.D. degrees from the University of
Idaho. He has worked in native fish conservation and fisheries management and research in the Pacific
Northwest for the past 32 years.
Michael K. Young is a Research Fisheries Biologist with the Rocky Mountain Research Station in
­ issoula, MT. He holds a B.S. and M.S. from the University of Montana and a Ph.D. from the University
M
of Wyoming. His work focuses on the conservation biology of cutthroat trout.
Jason B. Dunham is an Aquatic Ecologist with the U.S. Geological Survey, Forest and Rangeland
­Ecosystem Science Center, in Corvallis, OR. He holds a B.S. from Oregon State University, an M.S. from
Arizona State University, and Ph.D. from the University of Nevada-Reno. His work focuses on ­native
and nonnative species in freshwaters.
Cover photos and images clockwise from top: A gabion barrier constructed on North Barrett Creek in the Snowy
Range, Wyoming, to prevent upstream movement by nonnative trout (photo by G. T. Allison, U.S.D.A. Forest Service). Nonnative brook trout (top) and native greenback cutthroat trout (bottom) from Hidden Valley Creek, Rocky
Mountain National Park, Colorado (photo by K. D. Fausch, Colorado State University). Culvert beneath ­forest road
in Lost Creek, Idaho (photo by J. B. Dunham, U.S. Geological Survey).Map of the North Fork Coeur d’Alene River
basin, Idaho, showing stream segments where nonnative brook trout have invaded (red) and potential brook trout
habitat (yellow). See Figure 7 in Section VI for details.
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Executive Summary
cases environmental factors or isolating mechanisms apparently reduce risks of introgression. More research is needed to
understand where each species will invade and displace native
species versus where they will coexist or fail to invade.
Because the outcome of invasion and isolation can differ
among species of native salmonids, and among regions or
habitats, installing barriers is often a complex problem that must
be analyzed carefully. Four key questions provide a framework
for considering the use of barriers to prevent invasions and
conserve native fishes:
Native salmonid populations have declined throughout the
world due to a host of human influences, including habitat degradation and loss, invasion of nonnative fishes, and overfishing.
In many regions, suitable coldwater habitats for salmonids are
found mainly in protected natural areas, so that these fishes are
increasingly relegated to smaller and more isolated pieces of their
former native ranges. Moreover, populations of native trout and
charr in these relatively undisturbed habitats are often at further
risk from invasion by nonnative salmonids from downstream.
Faced with this dilemma, fisheries managers frequently consider
using barriers to upstream movement to prevent invasion and
displacement or hybridization. However, in doing so they face a
tradeoff because isolating native salmonid populations in small
headwater habitats may also increase their risk of extinction.
Here we focus on native salmonids in the inland western U.S.
(for example, cutthroat trout, Oncorhynchus clarkii; bull trout,
Salvelinus confluentus), but this is an important problem for
salmonids and other stream biota worldwide.
This monograph reviews the state of knowledge about the
factors that affect this tradeoff between invasion and isolation, and presents a framework for analyzing it and prioritizing
conservation actions. Barriers to prevent invasions can pose
problems for salmonids because these fish often need to move
to complete their life history. Although anadromous salmonids
are known for their extensive migrations, freshwater trout,
charr, grayling, and whitefish also show remarkable flexibility
in life history and diverse movement behaviors. Movements
allow fish living in patchy environments to maximize fitness by
placing each life history stage in habitats that provide optimum
growth and survival. Likewise, in environments that fluctuate
seasonally or among years, diverse life histories, and movements allow fish to avoid harsh conditions, or recolonize habitats
after catastrophes.
Isolation of fish populations using barriers can extirpate
mobile life history types, restrict fish populations to habitats
inadequate for long-term persistence, and prevent natural
recolonization after catastrophes. Both direct and indirect evidence from field research indicate that isolated populations of
cutthroat trout and bull trout are more likely to be extirpated in
smaller watersheds, but studies have been conducted in only a
few regions. These case studies suggest that these salmonids
need approximately 10 km of suitable stream habitat to persist
for 25 to 50 years and maintain genetic diversity, although the
long-term fate and evolutionary potential of populations in such
small watersheds is unknown. Much more research is needed
to refine these estimates for all taxa, because the amount of
habitat needed for persistence is likely to vary strongly with
climate and basin characteristics.
Invasion by three commonly introduced salmonids (brook
trout, Salvelinus fontinalis; brown trout, Salmo trutta; and rainbow trout, O. mykiss) can also extirpate inland native stream
salmonids, by competition and predation or introgressive hybridization. For example, field research on brook trout in the
inland western U.S. showed that they rapidly invade upstream
in some regions, and displace or completely eliminate native
cutthroat trout by reducing survival during the first two years of
life. However, in other regions brook trout invasions appear to
have stalled, perhaps limited by environmental factors. Likewise,
nonnative rainbow trout often hybridize with native cutthroat
trout, and nonnative brook trout with bull trout, but in certain
1. Is a native salmonid population of important conservation
value present?
2. Is the population vulnerable to invasion and displacement?
3. If the native salmonid population is isolated to prevent
invasion, will it persist?
4. If there are multiple populations of value, which ones are
priorities for conservation?
The first step in any conservation plan is to set clear management objectives by considering the values embodied in
a native salmonid population. Three conservation values
emerge from the literature: evolutionary, ecological, and socio-economic. Evolutionary value focuses on distinct species,
races, and populations, such as those protected in the U.S.
by the Endangered Species Act, many of which are adapted
to specific environments. For example, in some regions small
remnant populations of native cutthroat trout represent a rare
and significant genetic resource. Ecological value includes
important ecological processes and functions at the population,
community, and ecosystem levels. For example, species like
salmon and trout have strong effects on trophic webs in aquatic
systems, which may also extend into the terrestrial ecosystem.
Also of ecological value are species and populations that
are self-sustaining, resilient, adaptable, and require minimal
inputs of resources to maintain them. Socio-economic values
include recreational and economic benefits from fishing and
tourism. Although values may overlap, often it is not possible
to conserve all of them simultaneously in any one location, so
effective management will involve clearly defining priorities.
The second question to address for a population of conservation value is whether it is vulnerable to invasion and
displacement by, or hybridization with, a nonnative salmonid.
This depends on the ability of the invader to be transported to
the basin or spread to the target location, establish a reproducing population, and dominate in interactions with the native
salmonid. Currently, most transport of nonnative salmonids is
by unauthorized introductions by the public. Spread from these
initial invasions may be facilitated by long distance movements
(that is, “jump dispersal”), or hampered by barriers to upstream
movement like cascades and waterfalls. Headwater stocking
facilitates downstream dispersal throughout entire drainage
basins. Once nonnative salmonids arrive, they tend to establish populations where abiotic regimes of temperature, flow,
and natural disturbances allow reproduction and recruitment.
Nonnative salmonids may have impacts through competition,
predation, hybridization, or spreading parasites or pathogens.
Empirical evidence has demonstrated strong individual biotic
interactions (competition or predation) with native salmonids,
particularly during the first two years of life. Hybridization is
also commonplace and spreading in some cases, but rare or
localized in others. Recent evidence suggests that movement
of nonnative trout may transport nonnative parasites (such
i
as Myxobolus cerebralis, which causes whirling disease)
upstream to contact native salmonids.
The third question for valuable populations vulnerable to
displacement by a nonnative salmonid is whether the native
fish population will persist if isolated using a barrier to prevent
upstream invasion. Both direct and indirect field evidence indicate that larger interconnected patches sustain populations
longer than do smaller isolated ones. Natural disturbances
like fire, flood, drought, and freezing, as well as more subtle
climate changes that affect stream flow and temperature, can
drive small salmonid populations to extinction. Isolation can
eliminate migratory life histories that confer resilience through
large fecund adults, the opportunity for gene flow, and recolonization following catastrophic events. Isolation may also interact
with habitat degradation and overfishing to reduce population
persistence in small isolated stream segments.
The fourth question to address is which native salmonid
populations of value are deemed the highest priorities for
conservation, given limited management resources. Priorities
should be based on the relative conservation values of each
population, the risk of losing them, and the feasibility of reducing that risk through management. These values, risks, and
opportunities for management may vary strongly across any
area of interest, so a strategic, spatially explicit approach is
useful. It is important to understand the natural spatial structure
of native salmonid populations and their dynamics through
time, rather than arbitrarily focusing on individual streams or
other geopolitical units. Planning to conserve native salmonids
requires maintaining broad representation of the remaining ecological diversity, ensuring there is redundancy (more than one
population of each type) to buffer against loss, and considering
the resilience and resistance of populations to environmental
disturbances. An appropriate goal is to seek a set of populations
that represent the diversity of possible conservation values, and
that can interact, persist, and evolve to maintain those values
through time.
Two examples from the Coeur d’Alene River, Idaho and
the Little Snake River, Colorado-Wyoming, illustrate use of
the framework in markedly different situations. In the Coeur
d’Alene River, brook trout and rainbow trout have invaded
some streams, but large amounts of relatively undisturbed
habitat remain uninvaded. Here, there are opportunities to use
barriers judiciously to protect genetically pure populations of
westslope cutthroat trout (O. c. lewisi) with evolutionary values
from further invasion, and to sustain key ecological and socioeconomic values by either retaining or restoring passage for
migratory life history forms. In contrast, in the Little Snake River,
nonnative trout have invaded almost completely so that many
barriers have been installed in headwater streams to protect
the evolutionary legacy of remnant Colorado River cutthroat
trout (O. c. pleuriticus) populations. Unfortunately, there is little
possibility for these populations to support the full spectrum of
ecological and socio-economic values without large efforts to
install barriers and eradicate nonnative trout from downstream
segments. Because many of the remnant isolated populations
are small, persistence of native cutthroat trout in the basin may
depend on maintaining resilient populations through intensive
habitat protection or restoration, and on reintroductions by humans if they are extirpated. Priorities could focus on removing
nonnative salmonids to allow extending the larger populations
downstream and connecting them with other populations.
We developed a decision-space diagram to allow biologists managing a particular basin and set of native salmonid
populations of conservation value to consider the available
options and make strategic decisions under different circumstances like those in the examples above. These decisions
will be influenced primarily by: 1) the degree of isolation of the
populations of interest and 2) the degree of invasion threat. If
most remaining native populations are in headwater streams
and are clearly vulnerable to invasion and displacement, then
barriers will be required to protect them from rapid extinction.
Given this, strategic decisions available include the number,
size, quality, and spatial distribution of habitat patches; the
representation and redundancy of populations needed to conserve remaining diversity; and effective protocols for removing
nonnative salmonids, and translocating native salmonids to
found new populations. The goal should be to develop larger
and more connected habitat patches and to move the invasion
threat farther away, even though in many cases this is not
fully achievable at present. At the other end of the spectrum,
if a basin has many large interconnected patches of suitable
habitat with native salmonids and the invasion threat is distant
or weak, then strategic decisions will involve preventing invasions by controlling the sources and monitoring their spread,
preventing fish movement barriers and other types of habitat
fragmentation, and maintaining natural ecological processes
that provide complex habitats. Other regions of the decision
space require different strategies for management, or a mix of
strategies, and additional strategic decisions can be made about
barrier placement or removal under different scenarios.
Finally, we emphasize that the appropriate use of barriers
will depend not only on the current threats of invasion and
isolation, but also may change through time as human values
shift, as anthropogenic disturbances change, and as climate
and natural disturbances are altered. Because we don’t fully
understand the risks associated with barriers and invasion of
non-native salmonids, most management should be considered an experiment and carefully evaluated to improve future
efforts. Effective management will require careful monitoring
of populations, evaluating the results of such management
experiments (for example, the installation or removal of barriers), and investigating the basic processes of invasions and
population persistence.
ii
Contents_____________________________________________
Page
Executive Summary...............................................................................................................i
Introduction............................................................................................................................1
How to Use this Publication...................................................................................................1
Section I: An Important Problem............................................................................................2
Section II: Movement and Salmonid Life Histories................................................................4
Section III: Effects of Isolation...............................................................................................5
Direct Evidence................................................................................................................6
Indirect Evidence.............................................................................................................8
Section IV: Effects of Invasions by Nonnative Salmonids.....................................................9
Displacement of Native Salmonids..................................................................................9
Hybridization, Pathogens, and Parasites.......................................................................11
Section V: A Framework for Analyzing Tradeoffs.................................................................12
Conservation Values......................................................................................................12
Vulnerability to Invasion.................................................................................................16
Isolation and Population Persistence.............................................................................19
Considering Priorities among Multiple Populations.......................................................21
Section VI: Two Examples of Invasion and Isolation Tradeoffs...........................................24
Coeur D’Alene River......................................................................................................25
Little Snake River...........................................................................................................31
Section VII: Making Strategic Decisions..............................................................................34
Epilogue: What do We Need to Know?...............................................................................36
References..........................................................................................................................37
Acknowledgments________________________________________
We thank K. Crooks, D. Simberloff, B. Nehring, A. Suarez, B. Noon, D. Peterson, R. Knight, and
J. Savidge for information and discussion on the invasion-isolation tradeoff, and S. Dekome, E. Lider,
M. Davis, J. Dupont, and N. Horner for providing information and review for the Coeur d’Alene River
basin example. S. Barndt, M. Enk, A. Harig, F. Rahel, B. Riggers, K. Rogers, and B. Shepard provided
constructive reviews that helped improve the manuscript. We also thank Kate Walker of Region 1 of
the USDA Forest Service for help securing funding, Dona Horan for preparing the maps, K. Morita,
G. Allison, R. Stuber, R. Thurow, and P. Valcarce for providing images, and R. Dritz for checking
references. This synthesis was prepared during a sabbatical by K. Fausch at the Boise Forestry
Sciences laboratory of the Rocky Mountain Research Station, which was generously supported
by the USDA Forest Service (RJVA 04-JV-11222014-175, administered by B. Rieman), Trout
­Unlimited (administered by W. Fosburgh), and Colorado State University. Finally, we are indebted
to the many fish biologists and ecologists whose dedicated research over many years provided
the information needed to develop this synthesis.
iii
Strategies for Conserving Native
Salmonid Populations at Risk _
From Nonnative Fish Invasions:
Tradeoffs in Using Barriers to Upstream Movement
Kurt D. Fausch_
Bruce E. Rieman_
Michael K. Young_
Jason B. Dunham
Introduction_____________________
Native salmonid populations have declined throughout
the world due to a host of human effects, including habitat
degradation and loss, invasion of nonnative fishes, and
overfishing. In many regions, suitable coldwater habitats
for salmonids are found mainly in protected natural
areas, so that these fishes are increasingly relegated to
small pieces of their former native ranges. Moreover,
populations of native trout and charr in these relatively
undisturbed habitats are often at further risk from invasion by nonnative salmonids, usually from downstream.
Faced with this dilemma, fisheries managers frequently
consider using barriers to upstream movement to prevent
invasion and displacement. However, in doing so they
face a tradeoff because isolating native salmonid populations in small headwater habitats may also increase
their risk of extinction.
This monograph describes this widespread problem
for stream salmonids, reviews the state of knowledge
about the factors that affect extinction risks from invasion
versus isolation, and presents a framework for analyzing
this tradeoff and prioritizing conservation actions. We
first briefly review the natural population structure and
life history of stream salmonids, focusing on movements
that link populations across habitats dispersed throughout
the riverscape. We then discuss the effects of isolating
salmonid populations above barriers, which restricts
these movements and may drive populations extinct,
and the effects of invading nonnative salmonids, which
may move upstream (or downstream in some cases) and
displace native salmonids.
Second, we present a conceptual framework to guide
analysis of the tradeoff, focusing on four main questions
concerning conservation value, vulnerability to invasion,
persistence if isolated, and priorities when conserv-
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
ing multiple populations. We provide two examples to
­illustrate the use of the framework and show the key
elements in the decision process, and end by discussing
opportunities for making strategic decisions when faced
with this tradeoff.
This paper is limited in scope to native and non-native
salmonids and the stream networks they inhabit or could
invade, because salmonids dominate vertebrate communities in coldwater streams and are often the main focus
of management. However, the tradeoff is also relevant to
other aquatic taxa, both vertebrate and invertebrate, so
the framework we propose may serve as a foundation for
considering similar issues for these groups.
How to Use this Publication________
This report is written in seven major sections, to allow
readers with different backgrounds to find information
quickly. Some points are supplemented with additional
material that appears in text boxes, and some important
concepts are presented first in the review of knowledge
and used again in the framework for analyzing the
tradeoff. Sections are arranged as follows:
• Section I describes why the tradeoff between invasion and isolation is an important problem for
salmonids worldwide.
• Section II summarizes the life history and population structure of stream salmonids, and describes
why movement may be critical for their persistence
and long-term viability.
• Section III presents what is known about the effects
of isolation by movement barriers on persistence of
stream salmonid populations.
• Section IV discusses the effects of invasion by
nonnative salmonids on native salmonids.
• Section V presents a conceptual framework for
analyzing the tradeoff between invasion and
­isolation.
• Section VI includes two examples with different
constraints to illustrate the use of the framework and
show the key elements in the decision process.
• Section VII discusses opportunities for making
strategic decisions when conserving sets of native
trout populations in specific watersheds, given different degrees of isolation and invasion threat.
Section I: An Important Problem____
Habitat degradation and loss, invasion by nonnative
salmonids, and overfishing have reduced the distribution of native salmonids in many regions throughout
the world to small enclaves of headwater habitat where
they are vulnerable to extinction (Rieman and others
1997b, Harig and others 2000, Kitano 2004). For example, in many watersheds of the eastern U.S., from
New England to the southern Appalachian Mountains,
1
native brook trout (Salvelinus fontinalis) populations
are now restricted to headwater streams that are fragmented by dams, heavily sedimented, and close to dense
human populations (Hudy and others 2004). These
same watersheds simultaneously have been invaded by
nonnative rainbow trout (Oncorhynchus mykiss) and
brown trout (Salmo trutta) that displace brook trout from
downstream habitats (Larson and Moore 1985). Similar
combinations of habitat fragmentation and nonnative fish
invasions endanger native salmonids and other stream
organisms in other regions worldwide (see Box 1). These
examples illustrate that, although nonnative salmonids
do not invade everywhere they are introduced (Fausch
and others 2001, Adams and others 2002, Dunham
and others 2002a), three widespread invaders (brook,
brown, and rainbow trout) have established reproducing
populations in many regions outside their native ranges
(for example, see Lever 1996, Thurow and others 1997,
Fuller and others 1999, Cambray 2003b), and often
combine with habitat loss to limit native salmonids or
other native fishes to small headwater streams above
movement barriers.
Box 1. Examples of habitat fragmentation and
invasion for stream fishes worldwide.
•Native charr in northern Japan - Habitats of
native Dolly Varden (Salvelinus malma) and whitespotted charr (Salvelinus leucomaenis) in Hokkaido
Island, northern Japan, have been highly fragmented
by channelization and erosion control dams constructed mainly during the last 35 years (Morita
and Yamamoto 2002), and downstream populations
are now being displaced by nonnative rainbow trout
and brown trout (Takami and Aoyama 1999, Kitano
2004).
•Native brown trout in Switzerland - Habitat for
native brown trout in tributaries to the Rhine River
has been fragmented by human activities over a long
period. These populations now face invasion by
rainbow trout, which are apparently better adapted
to the altered flow regime (Peter and others 1998).
•Native stream fishes in the Southern Hemisphere
- Native stream fishes in New Zealand, Australia,
Venezuela, Chile, Sri Lanka, and South Africa are
being decimated by habitat alteration and introduced
brown trout and rainbow trout, and persist primarily
in small headwater enclaves above waterfalls that
the nonnatives cannot ascend (Townsend and Crowl
1991, Crowl and others 1992, Pethiyagoda 1994, Ruiz
and Berra 1994, Pefaur and Sierra 1998, Pascaul and
others 2002, Cambray 2003a, Gajardo and Laikre
2003).
•Other native stream biota - Native stream fishes
are blocked from Great Lakes tributaries by barriers
built to exclude nonnative sea lamprey (Petromyzon
marinus; Smith and Jones 2005), and certain life
stages of native amphibians and invertebrates that
use stream corridors in the Rocky Mountains are
also blocked by barriers (Clarkin and others 2003,
Adams and others 2005).
1 Fishes of the genus Salvelinus are charr, but here we use the accepted common names for brook trout and bull trout (S. confluentus). Throughout, we use the common names of fishes published
by the American Fisheries Society (Nelson and others 2004).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
Dams, diversions, and culverts that create barriers to
upstream movement are among the most common agents
that fragment habitat. In the conterminous U.S., there
are estimated to be about 77,000 dams at least 2 m high
(6 feet) that store at least 62 X 103 m3 (50 acre-feet) of
water (U.S. Army Corps of Engineers 2006). This total
does not include the huge number of smaller dams and
diversions for domestic water supply or micro-hydropower developments, nor the hundreds of thousands of
road culverts that create impassable barriers for fish
(GAO 2001). In addition, after early settlers degraded
downstream habitats and overexploited fish populations
(Wohl 2001), they often translocated fish above natural
barriers like waterfalls to establish new populations in
headwater habitats. Regardless of whether barriers to
upstream movement are anthropogenic or natural, they
fragment stream habitat in watersheds, simultaneously
protecting native salmonid populations from impending invasions but potentially hastening their extinction
from environmental catastrophes or small population
effects (Caughley 1994). They also disrupt movements
of other fishes, amphibians, and invertebrates (Warren
and Pardew 1998, Vaughan 2002, Clarkin and others
2003, Adams and others 2005), and may have similar
effects on these taxa.
In this monograph, we focus on the tradeoff in extinction risks for native salmonids of the inland western
U.S. Native cutthroat trout (O. clarkii) were restricted
to headwater streams by early mining, logging, grazing, overfishing, and water diversions (Gresswell 1988,
Young 1995), and many of these remnant habitats were
subsequently invaded by nonnative brook, brown, and
rainbow trout that exclude or hybridize with cutthroat
trout (Figure 1; Behnke 1992, Dunham and others
Figure 1—Populations of native cutthroat trout
(top; S. Emmons image) in the inland western
U.S. have often been invaded by nonnative
brook trout (bottom; K. Morita image) originally
introduced in either downstream or upstream
habitats.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
2002a). Likewise, bull trout and native rainbow trout in
the Pacific Northwest and Canada face similar problems
(Rieman and others 1997b, Thurow and others 1997).
Although this tradeoff is just coming to light in many
regions of the world, it has become a controversial topic
in the inland western U.S. There, fish biologists that
seek to conserve native cutthroat trout and bull trout
disagree over the merits of building new barriers to
limit upstream invasions and preserve remaining native
salmonid populations versus removing barriers to allow
movement of native fish and thereby enhance the chances
for population persistence (Dunham and others 2002a,
Peterson and others 2004, Shepard and others 2005).
This issue is also important because conservation biologists in general face the same tradeoff – that managing
ecosystems to address habitat fragmentation precludes
isolating them to prevent invasions. Some conservation
biologists have proposed connecting fragments of habitat with corridors to reduce extinction risks (Beier and
Noss 1998, Dobson and others 1999), whereas others
have argued that increasing connectivity opens the door
to nonnative species and diseases (Simberloff and Cox
1987, Simberloff and others 1992, Hess 1994). However,
there are few well-documented cases of the problem
or potential solutions, even in the extensive literature
on landscape connectivity (Bennett 1999, Soulé and
Terborgh 1999, Crooks and Suarez in press).
Section II: Movement and Salmonid
Life Histories____________________
Barriers and invasions potentially pose problems for
salmonids because these fishes often move to fulfill their
life history. Within a fish’s lifetime, local environments
often become marginal or inadequate for its needs,
prompting movement to increase or maintain fitness
(Thorpe 1994, Brannon and others 2004; see Box 2). A
general model of life history developed for fishes (Harden
Jones 1968), and refined for stream fishes (Schlosser and
Angermeier 1995) and salmonids (Northcote 1997),
holds that 1) movements link together habitats required
for spawning, growth, and refuge from harsh conditions,
and 2) these habitats are often dispersed throughout the
riverscape (Schlosser 1991, Fausch and others 2002).
Although anadromous salmonids are well known for
their spectacular migrations to and from the ocean,
freshwater salmonids also show a great diversity of
movement behaviors (for example, Meka and others
2003; see Box 2). Salmonids show remarkable flexibility
in expression of life-history types, including co-occurring (sympatric) populations of the same species with
Box 2. Fish movement and migration.
Theory: Movement of organisms, both within and
between generations, is a response to an environment
that is patchy in space or through time (Wiens 1976).
The theory of animal migration is based on the premise
that organisms move to place each life-history stage
in suitable habitat to maximize fitness (Northcote
1978, 1992, Dingle 1996, Hendry and others 2004a,
2004b).
Migratory Life History Types: Fishes that migrate
within freshwater are termed potamodromous
(Gresswell 1997), but four additional terms are more
commonly used to describe the diversity of movement
behavior in freshwater fish (see below). Nevertheless,
salmonids display great plasiticity and variation in
movement behavior, which often defies this simple
typology.
Lacustrine salmonids complete their life cycle
entirely within lakes.
Adfluvial fish migrate from lakes into rivers to
spawn, and juveniles later descend to lakes to grow.
For example, “coaster” brook trout in tributaries of
Lake Superior (Huckins and others in press), and
Lahontan cutthroat trout in Pyramid Lake, Nevada
(Snyder 1917, Behnke 1992), display adfluvial life
histories, spawning in tributaries but moving downstream to large lakes to grow to large adult sizes.
Fluvial fish complete their life cycle within riverine
environments, but may use tributaries for spawning
and juvenile rearing. For example, bull trout in the
northern Rocky Mountains and Dolly Varden in rivers in northern Japan display fluvial life histories,
spawning and rearing in tributary streams but moving
downstream into larger rivers to grow to adulthood
(Rieman and McIntyre 1993, Rieman and Dunham
2000, Koizumi and Maekawa 2004).
Resident salmonids complete their life cycle entirely
within tributary streams, but may range widely seeking suitable habitats for spawning, feeding, or refuges
(Gowan and others 1994, Fausch and Young 1995).
different life ­histories, alternative life histories within
a common gene pool, and phenotypic plasticity of a
single genotype under different environmental conditions
(Hendry and others 2004a, Hutchings 2004). Distinct
forms that co-occur in the same watershed, or even the
same stream, allow a single species to exploit diverse
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environments (Carl and Healey 1984, Bagliniere and
others 1989, Brannon and others 2004) or persist in the
face of variable or uncertain ones (Rieman and Clayton
1997). Populations may include both resident and migratory individuals (Rieman and others 1994, Zimmerman
and Reeves 2000), and the migratory forms may display
substantial variation in the extent and timing of movements (Brannon and others 2004). The expression of
different life-history types is apparently partly under
genetic control, yet plastic in response to environmental
conditions (Gislason and others 1999, Hendry and others 2004a), which can produce strong variation in life
histories both within and among distinct environments
(Taylor 1990, Jonsson and Jonsson 1993, Yamamoto and
others 1999). Recent evidence also suggests that when a
species is exposed to a novel or changing environment,
new life histories (for example, resident vs. migratory
forms, seasonal timing of migration) may emerge relatively quickly, in several decades or less (Näslund and
others 1993, Quinn and others 2001, Hendry and others
2004b, Riva Rossi and others 2004).
Movement is important not only in spatially patchy
environments, but also in those that fluctuate through
time so that fish are periodically reduced or extirpated in
some habitat patches. Dispersal that contributes to gene
flow, provides demographic support (that is, immigration
that offsets mortality or increases population growth
rate; see Stacey and Taper 1992, Hilderbrand 2003),
or allows recolonization can be critical to persistence
of these populations across generations. Subdivided
populations that interact strongly through movements
among them are termed metapopulations (Hanski 1999),
although salmonid populations are unlikely to fit simple
theoretical models (see Harrison and Taylor 1997, Dunham and Rieman 1999). Such metapopulation dynamics
may be particularly important in watersheds where major
disturbances like drought, flood, fire, and debris flows
create and destroy habitat, leaving a mosaic of habitats
in different successional stages (Reeves and others
1995, Neville and others in press). Salmonid populations that have evolved in these dynamic environments
may also have evolved diverse life histories, especially
movement behaviors, that ensure long-term persistence
at larger scales (Thorpe 1994, Heino and Hanski 2001).
For example, migratory fish may be critical to long-term
persistence because returning adults that fail to home to
their natal watershed disperse and recolonize or support
other habitats (Quinn 1993, Tallman and Healey 1994,
Hendry and others 2004b). In other cases diversity in
migratory timing and the occurrence of multiple age
classes “at large” in downstream rivers, lakes, or oceans
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during major disturbances can allow fish to quickly
recolonize the disrupted habitats (Rieman and others
1997a) or move to newly created ones (Leider 1989).
Small-stream salmonids may not exhibit the dramatic
migrations of forms with access to larger habitats, but
movement is still a critical behavioral component.
These fishes may move tens to thousands of meters
daily, seasonally, or during their lifetime to seek suitable habitats. For example, foraging cutthroat trout in
small streams have been observed to move over 150
m during the diel cycle (Young and others 1997), and
spawning runs of such fish have exceeded 2 km (Young
1996). Movements in fall and winter, associated with
declining water temperatures and ice formation, have
been of similar magnitude (Brown and Mackay 1995,
Jakober and others 1998). Moreover, mobility is common
among cutthroat trout of all ages in small streams, and
is not simply attributable to a few wandering individuals
(Peterson and Fausch 2003a, Schmetterling and Adams
2004).
Thus, movement plays important roles in stream salmonid populations, and yet remains poorly described
(Gowan and others 1994, Rieman and Dunham 2000).
The advent of advanced technologies such as radiotelemetry, PIT tags, and otolith microchemistry has enabled
recognition of some aspects of trout movements (Rieman
and others 1994, Fausch and Young 1995, Meka and
others 2003, Wells and others 2003), but a full understanding awaits sampling that is continuous over scales
of space and time broad enough to track individuals of all
life histories throughout their life cycles (Baxter 2002,
Fausch and others 2002). Such understanding remains a
central issue for research because movement may be the
key to population persistence in variable and changing
environments (Rieman and Dunham 2000).
Section III: Effects of Isolation______
Given the potential for extensive movements by salmonids both within and among generations, it is clear that
barriers like dams or culverts could extirpate mobile life
history types, restrict populations to habitats inadequate
for long-term persistence, and prevent recolonization
in the event of population extinction. Barriers might
also disrupt movement of other aquatic organisms, and
potentially change the structure of whole communities. However, even for salmonids, which are relatively
well studied, evidence for negative effects of isolation
in headwater streams is still emerging. There is some
direct evidence from studies of barriers in Japan and
translocation of cutthroat trout in the inland western
U.S., and indirect evidence from patch occupancy in
both countries. There is also indirect evidence from
molecular genetics and population models, which we
summarize briefly.
Direct Evidence
Direct evidence for negative effects of isolation on
headwater stream salmonid populations comes from two
studies where barriers were built, or fish were translocated
above barriers, and the results measured. In both cases,
investigators fit relationships between watershed area and
species occurrence, allowing empirical estimates of the
watershed areas required for a specified probability of
population persistence (for example, 0.50 or 0.90). We
caution that these estimates may not be easily generalized to other regions, because watersheds that support
habitat of sufficient size, length, or complexity to sustain
native salmonid populations can differ markedly in
area, depending on climate, geology, and other factors.
In addition, different species, and even subspecies, of
salmonids may have different life history characteristics that allow persistence despite isolation. Despite
this complexity, these studies do show that watershed
area is strongly associated with population persistence,
and they provide first approximations of relevant size
thresholds.
Morita and Yamamoto (2002) measured occurrence
of whitespotted charr in 52 headwater fragments above
erosion control dams (all >2 m high) in mountain streams
of southern Hokkaido Island, Japan, and found that both
watershed area and time since isolation had important
effects on population persistence (Figure 2). The authors
inferred that all sites had charr populations before the
dams were built, because habitat was no different than
at 32 control sites that supported charr. Seventeen of
the 52 populations had been extirpated in the 30 years
since most barriers had been built, and 12 of the 35
remaining populations were predicted to go extinct in
the next 50 years. Overall, they predicted that a minimum watershed area of 2.3 km2 is needed to have a 50
percent chance of sustaining a population for 50 years,
and their figure shows that about 9 km2 is needed to
have a 90 percent chance. Southern Hokkaido receives
much more precipitation than the inland western U.S.
so much smaller watersheds sustain perennial streams.
However, this study shows that useful relationships
can be developed from historical records and relatively
simple field measurements.
Further work on extant populations of these charr
revealed additional ecological effects of the barriers.
Morita and others (2000) reported that growth rates of
whitespotted charr juveniles in three rivers were higher
above dams than below, likely due to the lower spawning and fry density upstream. This resulted in more fish
remaining resident than smolting, a common response to
increased juvenile growth rates (Thorpe and Metcalfe
1998, Hendry and others 2004a). Moreover, growth
rates of fish transplanted from below to above the dam
in one of the streams also increased and smoltification
decreased, indicating that the changes were due to phenotypic plasticity rather than rapid evolution.
In the second study, Harig and Fausch (2002) measured persistence of allopatric greenback cutthroat trout
(O. c. stomias) and Rio Grande cutthroat trout (O. c.
virginalis) in segments of 28 streams (1.0 to 20.5 km
long) in Colorado and New Mexico where they had been
translocated above barriers during the previous 3 to 31
years. The models they fit predicted that translocations
in watersheds >14.7 km2 had a >50 percent chance of
establishing a reproducing population, whereas in smaller
watersheds populations were more likely to die out or
remain at low abundance. Watersheds of >33 km 2 were
predicted to have a >90 percent chance of establishing a
reproducing population, a goal more in line with many
conservation assessments. The empirical data used to
develop the model showed that the 13 stream segments
where cutthroat trout reproduced consistently were, on
average, 7.3 km long (range: 1.8 to 20.5) and drained
watersheds that averaged 23 km2, although 6 of the
shortest also included lakes.
Virtually all of the historical stream populations of
these two subspecies of cutthroat trout were isolated from
nonnative salmonid invasions above natural barriers. In
most cases, they were probably translocated there by
early settlers. This suggests that at least some populations of these subspecies can adapt to such habitats and
persist for decades. Indeed, Harig and Fausch (2002)
found that 32 of the 70 historical populations persisted
in stream fragments with watersheds smaller than 14.7
km2. However, these may be at greater risk of eventual
extinction due to insufficient habitat, an effect Tilman
and others (1994) referred to as an “extinction debt” that
is yet to be paid (see Hanski 1997 for review). Finally,
we note that other studies have also reported loss of
stream fish populations above barriers, and reductions
in fish species richness upstream versus downstream of
barriers, in other temperate and tropical rivers (Winston
and others 1991, Fausch and Bestgen 1997, Peter 1998,
Schmutz and Jungwirth 1999, Pringle and others 2000).
These observations lend further direct evidence for the
negative effects of isolation.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
Figure 2—Probability of whitespotted charr (top; K. Morita image) persistence above erosion control (sabo) dams (bottom;
K. Fausch image) in southern Hokkaido Island, Japan, as a function of watershed area (from Morita and Yamamoto 2002;
used with permission). Curves (center) were predicted from a logistic regression model of presence/absence as a function of
watershed area, time since damming, and stream gradient for 52 sites above dams. The family of curves in the figure predicts
the probability of population persistence in headwater streams that have been isolated for different time periods (given average
gradient). Watersheds >10 km2 have a high probability of sustaining these charr populations for 50 years, according to the
model, but different salmonid species in different environments may have markedly different relationships.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
Indirect Evidence
Several lines of indirect evidence also indicate that
small watershed size and isolation reduce persistence of
headwater salmonid populations. This evidence comes
primarily from studies of patch occupancy (patches are
defined as continuous networks of suitable stream habitat;
Dunham and others 2002b), but other indirect evidence
also comes from colonization after barrier removal, and
genetics and population models (see Box 3). In the first
study of patch occupancy for salmonids, Rieman and
McIntyre (1995) reported that bull trout in the Boise River
basin in Idaho had approximately a 50 percent chance
of occurrence in patches >25 km2 (which support about
5 km of stream habitat), and more than an 80 percent
chance in patches >80 km2 (about 9 km of stream habitat).
This suggested that populations in smaller patches were
more vulnerable to local extinction than those in larger
patches. In a second study, Dunham and others (1997)
reported that Lahontan cutthroat trout (O. c. henshawi)
Box 3. Additional indirect evidence on isolation.
Genetics: Isolation of small populations can cause loss of genetic diversity
and random genetic drift (Rieman and Allendorf 2001). In turn, this can lead
to a loss of phenotypic variation and evolutionary potential (Allendorf and
Ryman 2002), and potentially to reduced effective population size (Waples
2004) and inbreeding depression, both of which could hasten extinction of
salmonid populations isolated above barriers (for example, Soulé and Mills
1998, Frankham 2005). However, these mechanisms have not been demonstrated in headwater stream salmonids, and there has been only limited work
to discriminate effects of habitat size and isolation. Examples include:
•Yamamoto and others (2004) found lower genetic diversity and increased
genetic drift in whitespotted charr above barriers versus below in three watersheds in Hokkaido.
•Wofford and others (2005) reported that genetic diversity decreased and
genetic drift increased for coastal cutthroat trout (O. c. clarki) isolated above
more versus fewer barriers in a small headwater stream in Oregon.
•Neville and others (in press) reviewed evidence for salmonids, and found
that barriers were consistently associated with lower genetic diversity within
populations.
Population models: Population models have been used to 1) estimate total
population sizes and compare them to theoretical guidelines, and 2) simulate
the effects of isolation, population size, survival, and environmental variation
on persistence and genetic variation.
• Rieman and Allendorf (2001) used individual-based models to estimate
population sizes needed to sustain genetic variation in bull trout. They ­a rgued
that few local populations would maintain genetic variation indefinitely
without gene flow from surrounding populations.
•Morita and Yokota (2002) used an individual-based model to predict
persistence of whitespotted charr populations. Persistence increased with
larger populations and higher adult survival, but decreased substantially for
small populations after 30 to 100 years in the model, suggesting that smaller
populations are at a higher risk of eventual extinction.
•Hilderbrand (2003) modeled inland cutthroat trout populations and found
that persistence increased with increases in population size, immigration
from adjacent source populations, and juvenile survival, and with decreases
in environmental variation.
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in the Great Basin were more likely to occur in patches
that were connected to other occupied patches along
stream networks. This suggested that populations were
more likely to persist in less isolated patches. Further
research in the Boise River basin showed that both larger
patch size and smaller distance to occupied patches were
important predictors of bull trout presence (Dunham and
Rieman 1999). In both locations, disturbances like floods,
droughts, fires, and debris flows can extirpate these two
salmonids from whole stream segments (Rieman and
Clayton 1997, Dunham and others 2002b), so movement
is required to recolonize habitat fragments and maintain
local populations.
Three recent studies have also shown the importance
of larger connected patches in supporting headwater
salmonid populations. Rich and others (2003) found
that bull trout were more likely to be present in tributaries of Montana watersheds where there was a strong
source population in the mainstem river downstream.
Koizumi and Maekawa (2004) reported that Dolly
Varden were more likely to occur in larger spring-fed
tributaries of a Hokkaido river that were closer to other
occupied tributaries than in smaller and more distant
ones. Young and others (2005) used data from extensive
electrofishing surveys to calculate that streams of 3.4 to
12.9 km in length were needed to sustain 500 to 5000
age-1 and older greenback or Colorado cutthroat trout
(O. c. pleuriticus). These population size thresholds have
been proposed as sufficient to maintain genetic diversity
(n=500) and long-term evolutionary potential (n=5000;
Allendorf and others 1997, Hilderbrand and Kershner
2000, Young and Harig 2001), rather than simply to
ensure relatively short-term persistence (that is, 50 to
100 years). It is typical that half or fewer of the extant
populations of native cutthroat trout in various regions
are found in streams long enough to support populations
of these sizes (see Kruse and others 2001, Young and
others 2002, Young and Guenther-Gloss 2004, Young
and others 2005). Nevertheless, indirect evidence from
field studies like these is particularly relevant because
they encompassed the large spatial scales at which land
managers must set priorities and make decisions about
resource management (Fausch and others 2002).
Once habitats are made available by removing barriers, recolonization of unoccupied habitat appears likely.
The many examples of barrier removal followed by rapid
and extensive upstream colonization in fishes (Hart and
others 2002 and citations therein) serve as circumstantial evidence that renewed access to headwater habitats
above dams, particularly spawning areas, may increase
the persistence of downstream populations (Smith and
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others 2000). Similarly, streams that have lost fish populations because of drought, severe floods, fire-related
disturbances, or pollution are often quickly repopulated
(Larimore and others 1959, Sigler and others 1983,
Fausch and Bramblett 1991, Rieman and Clayton 1997,
Roghair and others 2002), and habitats available for the
first time (for example, following glacial retreat; Milner
and Bailey 1989) typically develop fish populations in
a predictable sequence. In some circumstances, fishes
require extended periods to colonize newly available
habitats downstream (Hepworth and others 1997), or
may be more likely to colonize other habitats following
displacement by floods (Leider 1989) or in response
to high densities (Gowan and others 1994, Isaak and
Thurow 2006).
Section IV: Effects of Invasions by
Nonnative Salmonids_ ____________
The other main factor in the tradeoff addressed here
is upstream invasion by nonnative trout that may reduce
or eliminate native salmonids in headwater streams. As
we described above, rainbow, brown, and brook trout
have been transported and released in most countries
and regions that have suitable coldwater habitat for salmonids (Lever 1996, Rahel 2000). In the western U.S.,
many of these introductions began in the 1870s with
the arrival of large numbers of Euroamerican settlers,
who further spread these fishes by railroad, horseback,
and on foot (Wiltzius 1985, Alvord 1991, Wiley 2003).
State and federal agencies assumed responsibility for
most stocking thereafter (Wiltzius 1985), which still
continues in some parts of this region (Epifanio 2000).
For example, brook trout were first released in the 11
western states between 1872 and 1912 (MacCrimmon
and Campbell 1969) and are still stocked in 7 of them
(Dunham and others 2002a).
Displacement of Native Salmonids
Where barriers to upstream movement are lacking,
nonnative salmonids like brook trout can move rapidly
upstream and colonize habitat occupied by native salmonids (Gowan and Fausch 1996, Peterson and Fausch
2003a). There have been multiple observations of rapid
declines in native salmonid populations after these nonnative trout invasions, such as declines of cutthroat trout
after nonnative brook trout and rainbow trout invasion
in western U.S. streams (Gresswell 1988, Behnke 1992,
Young 1995). These effects have been studied in detail
using both experimental and correlative approaches.
At the scale of 1- to 10-km stream segments, several
field removal experiments in which all age classes of
the native and invaded populations were measured provide strong evidence of invader impacts. Peterson and
others (2004) found that abundance of native Colorado
River cutthroat trout fry and juveniles increased after
four years of brook trout removal in a Colorado stream,
compared to a control stream, because survival of age-0
cutthroat trout increased 13 fold and that of age-1 fish
doubled after brook trout removal. In contrast, no such
increase occurred in another pair of streams at higher
elevation, because cutthroat trout recruitment often
failed, apparently due to cold temperatures (Coleman
and Fausch 2006). Surprisingly, survival of adult cutthroat trout was unaffected by brook trout removal in
either set of streams.
Several other nonnative trout removal experiments
conducted at the stream segment scale have also been
followed by increased native trout abundance. Moore
and others (1983, 1986) removed nonnative rainbow
trout from entire lengths of four streams above barriers
in Great Smoky Mountains National Park for 6 years.
They reported increases in abundance of native brook
trout compared to three control streams where abundance changed little. Shepard and others (2002) found
that a westslope cutthroat trout (O. c. lewisi) population
increased seven fold after removing brook trout from
above a barrier in a Montana stream for 8 years. The
decade-long removal of brook trout with piscicides and
electrofishing from a stream in Crater Lake National
Park, Oregon resulted in an almost three fold increase
in abundance of native bull trout (Buktenica and others
2001, Renner 2005). However, encouraging anglers to
remove brook trout from an Alberta stream inhabited
by native bull trout and westslope cutthroat trout was
ineffective and failed to reduce brook trout numbers
(Stelfox and others 2001).
Strong declines in native trout after nonnative trout
invasion have led investigators to propose competition
and predation as likely mechanisms. These mechanisms
have been studied extensively in controlled laboratory
and field experiments at small scales (see reviews in
Fausch 1988, 1998, Dunham and others 2002a, Peterson and Fausch 2003b), but results are often difficult
to generalize to larger spatial and longer time scales
(Rieman and others 2006). An exception is experiments
designed to explicitly consider environmental factors like
temperature across a range of natural conditions, either
in the field (McHugh and Budy 2005) or by simulating
conditions in the laboratory (Taniguchi and Nakano
2000).
10
Finally, field sampling at various scales provides correlative evidence of the effects of nonnative trout invasions. For example, brook trout appear to have displaced
cutthroat trout in a Nevada watershed (Dunham and others 1999), and bull trout in an Idaho watershed ­(Rieman
and others 2006), where the native and nonnative species
co-occurred. However, the effects were highly variable
among streams. Rich and others (2003) showed that bull
trout occurrence in small low-gradient streams below
barriers in a Montana watershed was strongly associated
with complex habitat structure and strong neighboring
populations that could provide demographic support, but
was negatively associated with presence of brook trout
(Figure 3). Townsend and Crowl (1991) and McIntosh
(2000) reported strong negative correlations between
brown trout and native galaxiid abundance in stream
segments below barriers throughout two New Zealand
watersheds, and provided experimental evidence that
predation by large brown trout was the cause. At the
scale of channel units, Morita and others (2004) used
an extensive field survey to provide evidence that brown
trout excluded native whitespotted charr from individual
pool-riffle sequences in a Hokkaido stream, whereas
rainbow trout apparently excluded the charr from pools
to adjacent riffles. These data at relatively small spatial
scales support conclusions about the effects of these
nonnative trout invasions in Hokkaido derived from
broad field surveys (Takami and Aoyama 1999, Kitano
2004).
In contrast to examples where nonnative salmonids
clearly displaced native trout, other cases show that nonnative salmonids apparently failed to invade all habitat,
or had little effect on native trout when they did invade.
Fausch (1989) reported that brook trout failed to invade
certain high-gradient streams with cutthroat trout in
several Rocky Mountain regions (see also Rieman and
others 1999, Paul and Post 2001). He suggested various
hypotheses including limited upwelling groundwater
required for egg incubation and winter floods that scour
eggs or fry. Likewise, Adams and others (2002) found that
brook trout invasion had apparently stalled in more than
half the tributaries studied in a pristine Idaho watershed,
based on comparison with historical data (see Section V
for more details). Novinger and Rahel (2003) found no
increase in native cutthroat trout in four small Wyoming
streams isolated by barriers after 5 to 8 years of removing
brook trout by electrofishing, suggesting that the brook
trout had little effect. They attributed this primarily to
lack of critical resources or abiotic conditions needed
for cutthroat trout recruitment and survival (see Harig
and Fausch 2002, Sloat and others 2005, Coleman and
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
Figure 3—Bull trout presence in Montana
and Idaho watersheds is often positively
associated with complex physical habitat
and strong populations in neighboring
watersheds. However, bull trout (fish at
bottom) presence has also been found to
be negatively associated with brook trout
(top) in certain watersheds, suggesting that
this nonnative salmonid either uses different
habitats than bull trout or displaces them
(K. Morita image).
Fausch 2006), or to lack of demographic support from
adult spawners. Fausch and others (2001) found that
rainbow trout invasion in various regions worldwide
was most successful where flow regimes were similar to
those in their native range, and reported that this widely
introduced species was apparently unable to invade in at
least one region where flooding was frequent during fry
emergence. Although there has been no comprehensive
analysis to date of locations where nonnative salmonids
have failed to invade, and the environmental correlates,
such research is clearly needed to explain differences
in invasion success and provide tools for managing the
invasion-isolation tradeoff.
Hybridization, Pathogens, and
Parasites
Hybridization is an additional process by which nonnative salmonids can adversely affect native salmonids
(Allendorf and others 2001, 2004). Hybridization has
obvious genetic influences, but ecological effects are also
possible. Hybrids of many salmonid crosses are fertile
(Suzuki and Fukuda 1971), which leads to backcrossing
of hybrids with parental types and gene flow between
genetically distinct populations, subspecies, or species,
a condition called introgression (Rhymer and Simberloff
1996). If backcrossing is extensive and hybrid individuals
also mate among themselves, the resulting population
is termed a hybrid swarm. For example, introgressive
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hybridization of native cutthroat trout with rainbow trout
is widespread throughout the western U.S. (Leary and
others 1998, Peacock and Kirchoff 2004). Weigel and
others (2003) reported that hybridization of westslope
cutthroat trout with rainbow trout was more prevalent in
the downstream segments of an Idaho watershed, despite
an apparent absence of migration barriers. They inferred
that harsh abiotic factors upstream limited rainbow trout
more than cutthroat trout. In contrast, Hitt and others
(2003) reported that hybrids between the two species
were spreading upstream throughout a large Montana
watershed, and inferred that introgression would expand
because it appeared that hybridization was limited more
by demographic factors like movement than abiotic factors that could limit spread.
Other cases suggest that hybridization can be restricted.
For example, when brook trout invade and mate with
bull trout, hybrids and subsequent backcrosses appear to
have reduced fertility or survival (Kanda 1998, Kanda
and others 2002). Although such hybridization may not
result in hybrid swarms (Kanda and others 2002), it does
represent wasted reproductive effort by the native fish.
Because invading brook trout can mature at an earlier
age and have higher reproductive potential (Power 1980,
Kennedy and others 2003), hybridization could still lead
to displacement of bull trout (Leary and others 1993,
Rieman and others 2006). Recent evidence indicates
that maintenance of large migratory bull trout may
limit hybridization with smaller resident Dolly Varden
11
through size-assortative mating (Redenbach and Taylor
2003). Assortative mating and differential habitat use
(B. Rieman and J. Dunham, personal observation) might
similarly limit hybridization between large migratory
bull trout and smaller resident brook trout, further
highlighting the potential importance of migratory life
histories and stream interconnections in maintaining
the integrity of native salmonid populations.
Last, invading nonnative trout infected with pathogens
or parasites may transmit them to native salmonid populations. For example, the myxozoan parasite Myxobolus
cerebralis that causes whirling disease has a resistant spore
stage that can remain latent for decades (El-Matbouli
and others 1992), and is transmitted to trout fry through
oligochaete (Tubifex) intermediate hosts. Nehring and
Thompson (2003) and Nehring (2004) reported invasion of the parasite 25 km upstream in one Colorado
watershed, and 4 km upstream in two smaller streams,
apparently via rainbow trout and brown trout movement.
In the first case, this movement brought the parasite to
within about 3 km of a population of native Colorado
River cutthroat trout. In contrast, Peterson and others
(2004) found no evidence that brook trout invading native
cutthroat trout habitat in four other Colorado streams
were infected with the parasite, but the distance to an
infected source population was unknown.
Section V: A Framework for
Analyzing Tradeoffs_ _____________
The information presented above makes clear that
fisheries managers often face a dilemma: nonnative
trout invasions can reduce or extirpate native salmonid
populations, yet barriers to prevent these invasions can
lead to extinction of native salmonids because of isolation. However, the outcomes of invasions and isolation
differ among species and stocks of native and nonnative
salmonids due to their evolutionary history, and among
locations due to habitat characteristics, environmental
variation, and time since isolation. This means that installation or removal of barriers to promote persistence of
native salmonid populations is often a complex problem
that must be analyzed in the context of these ecological
and evolutionary constraints. Such efforts will also be
influenced by conservation goals and the available opportunities and management resources. Here, we propose
a conceptual framework that explicitly considers these
factors. We outline a process that incorporates the preceding synthesis of existing knowledge as well as the
experience of field biologists. We illustrate the concept
with examples from two basins in the Rocky Mountains
12
that have markedly different settings and constraints,
and end by considering an overall strategy for making
decisions.
Four key questions provide a framework for considering the use or removal of barriers to conserve native
fishes in any stream or network of streams.
1.Is a native salmonid population of important conservation value present?
2.Is the population vulnerable to invasion and displacement by nonnative salmonids?
3.If the native salmonid population is isolated, will
it persist?
4.If there are multiple populations of value, which
ones are priorities for conservation?
Although these questions are relatively simple, each
leads to more questions or dimensions that must be
considered in any decision (Table 1; see Dunham and
others 2002a), so we address each in turn.
Conservation Values
A first step in developing any native salmonid management plan is to consider the conservation values
associated with the populations of interest (Table 1).
Without a clear concept of the relative values of different
populations that could be managed, it is impossible to set
specific management objectives. What constitutes value
varies markedly among government and tribal agencies,
those who study or manage native fish species, and the
interested public. For natural resource management
agencies, conservation of biodiversity is a widely held
goal, and the conservation literature presents a diverse
set of guidelines toward that end (see Box 4).
We suggest that three general classes of conservation
value emerge from the literature (for example, Angermeier and others 1993): evolutionary, ecological, and
socio-economic (Figure 4). Evolutionary conservation
values are considered explicitly by the U.S. Endangered Species Act (ESA), which calls for designating
evolutionarily significant units and distinct population
segments (Waples 1995). Here the focus is on distinct
species, stocks, and populations that have evolved in,
and adapted to, unique and varied environments. Genetic and ecological divergence are important criteria
(Allendorf and others 1997) and the full representation
of diversity at regional scales (104 to 105 km2) has been
a focus of ESA-related management efforts to conserve
evolutionary legacies (Waples 1995, U.S. Fish and Wildlife Service 1998b). Some species such as westslope cutthroat trout and bull trout can show substantial genetic
and phenotypic divergence among tributaries within a
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Table 1—Four questions to consider when evaluating the tradeoff of managing ­native
salmonid populations using barriers to isolate them from invasions (after
­Dunham and others 2002a, S. Dunham unpublished). The answer to each
question depends on further dimensions or considerations (right column) for a
more complete analysis of the tradeoff.
Main questions
Further dimensions or considerations
1. Is a native salmonid population of important conservation value present?
• Evolutionary values
• Ecological values
• Socio-economic values
2. Is the population vulnerable to invasion and displacement by nonnative salmonids?
• Transport and spread
• Establishment
• Effects
oDisplacement
oCoexistence
oHybridization
oTransmit parasites or pathogens
3. If the native salmonid population is • “Small population” phenomena
isolated, will it persist?oLoss of genetic variability
oDemographic stochasticity
oEnvironmental stochasticity
oCatastrophes
• Loss of migratory life histories
• Synergistic factors
4. If there are multiple populations of value, which ones are priorities for conservation?
single river basin (102 to 103 km2), arguing for conserving
distinct populations at this scale as well (Allendorf and
Leary 1988, Leary and others 1998, Quinn and others
2001, Hendry and others 2004b). From this perspective,
evolutionary conservation values are highest for distinct
populations that represent a substantial or unique portion
of extant genetic diversity, occupy unique or remnant
habitats, or exhibit diverse or unusual life histories
(Allendorf and others 1997, Halupka and others 2003,
Shepard and others 2005).
Ecological conservation values include important
ecological processes and functions at the population,
community, or ecosystem levels of organization. For
example, species like salmon and trout can have strong
effects on trophic networks in aquatic systems (Simon
and Townsend 2003), which can even cross the ecosystem
boundary into adjacent terrestrial ecosystems (Spencer
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• Identify conservation units
• Prioritize populations
oRepresentation
oRedundancy
oPersistence
oFeasibility
• Consider viability
• Confront uncertainty
and others 1991, Willson and Halupka 1995, Baxter and
others 2004, Koyama and others 2005). As already discussed, some populations may be the key to persistence
of others through dispersal, gene flow, demographic
support, and metapopulation-like processes (Schlosser
and Angermeier 1995, Rieman and Dunham 2000).
Thus, dispersal from stronghold or source populations
may allow the entire network to persist in the face of
disturbance or changing environments. Similarly, the full
expression of life-history diversity within populations
may confer resilience to environmental disturbance (Rieman and Clayton 1997, Brannon and others 2004) and
in some cases resistance to invasion. Ecological value
can be far more varied than these simple examples, but
a key point is that the structure and function — that is,
the composition and roles of individuals or populations
— are ecologically important. We may value essential
13
Box 4. Components of value for biodiversity.
Different authors have proposed different systems for
considering the requirements for, and values of, sustaining biodiversity:
•Noss (1990) and Groves (2003) proposed that sustaining biodiversity requires conserving three attributes of
ecosystems: composition, structure, and function. Composition includes the unique elements such as species or
genes. Structure represents the distribution, interconnections, and interactions of those elements. Function
represents the key processes or ecological products or
influences that result (for example, secondary production).
•Noss and Cooperider (1994) described four explicit
values of biodiversity: direct utilitarian (for example,
harvestable protein), indirect utilitarian (for example,
clean water), recreational/aesthetic (for example, fishing), and intrinsic spiritual/ethical aspects.
•Allendorf and others (1997) focused on biological
values defined by the potential evolutionary and ecological consequences of extinction. Thus, extinction
may result in loss of evolutionarily distinct, rare, and
genetically diverse populations, as well as loss of
members of unique communities or environments, or
of “conservation umbrella” species.
•Callicot and Mumford (1997) outlined a classic debate in natural resource and biodiversity
management: whether ecosystem sustainability might
be met through either of two potentially distinct objectives - conserving ecosystem integrity or conserving
ecosystem health. Integrity refers to the native species
complex, interactions, and diversity, whereas health
describes a system that still retains its function and
organization but not necessarily the native elements.
The differences between the two have proven controversial. Although we can conserve ecological processes
without the native community, functional ecosystems
might still be substantially impoverished (Callicott
1995, Groves 2003). Nevertheless, conservation of
ecological process remains an important challenge that
should not be overlooked, even when there is little hope
of conserving native diversity or integrity represented
by “ghost” or “relict” species (Meyer 2004).
Figure 4—Distribution of three conservation values in
salmonid populations in natural and disrupted stream
systems. In natural systems (top) all values may be
represented in the same species distributions. In
systems disrupted by human development and the
introduction and invasion of non-native salmonids
(bottom), all of these conservation values often cannot
be supported in the same streams or fish communities.
For example, isolation of genetically pure native trout
populations may conserve primarily evolutionary
values (solid line), but limited ecological functions and
values (dashed line). There is currently much debate
about whether hybrid trout retain significant ecological
functions and values, but most are considered to
have limited or no evolutionary value. Hatchery trout
in a put-and-take fishery could have significant socioeconomic value but no evolutionary or ecological values.
This simple classification is not mutually exclusive
(for example, isolated pure trout may also have socioeconomic conservation values) and different values
may overlap depending on the system and the species
and forms involved. However, in disrupted systems
the potential to represent all conservation values
simultaneously is likely to be far more limited than in
natural systems.
14
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ecological services that are provided, or the ability of an
ecosystem and its elements to organize, reorganize, and
respond to changes by rebounding, adapting, and evolving. From a management perspective, we value species,
populations, and communities that are self-sustaining,
resilient, and adaptable because they don’t require continual input of resources and energy to maintain them.
However, a key issue is that although evolutionary and
ecological values sometimes overlap, they are not the
same thing (Callicott 1995, Callicott and Mumford 1997).
For example, a genetically introgressed cutthroat trout
population could retain significant ecological value, but
have little evolutionary value (Figure 4; Allendorf and
others 2004). From this perspective, ecological conservation values are higher for populations that are large,
productive, and interconnected, and fill roles similar to
those in the original ecosystem. Likewise, they are less
for populations that are vulnerable to extinction without
extraordinary conservation effort, or those that threaten
the persistence of ecosystem structure and function that
we hope to maintain (for example, fish populations that
are vectors of disease).
Socio-economic conservation values are perhaps the
most evident to fisheries managers who are anxious to
conserve the recreational and economic benefits from
fishing, tourism, and associated activities. These values
may be more obvious because they can be quantified
in economic terms, capture the public interest, and
often dictate the budgets and activities of management
agencies. They are also important because they can
leverage conservation resources for other species and
populations that may hold significant evolutionary and
ecological values. For example, conserving salmon may
also protect lamprey (Petromyzontidae; Lee and others
1997, Thurow and others 1997). Like the first two values, socio-economic values can also vary among local
populations of conservation interest.
Recognizing distinct evolutionary, ecological, and socio-economic conservation values for a particular population or set of populations is important because they may
afford very different opportunities for conservation and
the implementation or removal of barriers. In some large
wilderness rivers such as the Middle Fork Salmon River
in Idaho or the South Fork Flathead River in Montana,
undeveloped watersheds and native species complexes
clearly support substantial values in all three categories
(Figure 5). But habitat and populations in most river
systems throughout the region have been substantially
altered, so that convergence of all three conservation
values in any single population is rare (Figure 4). For
example, main-stem populations of cutthroat trout may
be large and composed of individuals with mobile life
Figure 5—In relatively pristine wilder­
ness river basins like the Middle Fork
Salmon River, Idaho, shown here
(Sulphur Creek, Frank Church River
of No Return Wilderness; R. Thurow
image), native salmonid assemblages
can support substantial evolutionary,
ecological, and socio-­economic values
simultaneously. However, most river
systems have been substantially altered,
so convergence of all three conservation
values in any single location is rare.
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15
histories, but tend to be introgressed and co-occur with
nonnative salmonids, whereas headwater populations are
generally small and isolated yet offer the best prospects
for being genetically pure. As suggested above, the mainstem populations may hold significant ecological value,
but little evolutionary value, whereas the opposite may
be true of the headwater populations. In these cases,
the prioritization of conservation efforts is challenging
because the tradeoffs are difficult or uncertain. Should
we favor populations that represent remnant elements
of evolutionary history or those that have the potential
to evolve and adapt in the future (Meyer 2004)? Should
we favor locally hybridized populations that may retain
some elements of local adaptation (Dowling and Childs
1992, Peacock and Kirchoff 2004), or should we replace
them with a genetically pure stock that represents many
pooled populations (Leary and others 1998)? Should
we maintain access for large migratory fish that support a key fishery even though local populations will be
vulnerable to invasion and hybridization, or should we
isolate the local populations to conserve genetic purity
(Allendorf and others 2004)?
It is clear that all three conservation values are important depending on the context, and often fisheries
managers will be faced with a mix of objectives and
possibilities. Opportunities to conserve evolutionary
values, for example, may be very restricted and viewed
as immediate priorities in a few distinct areas if they
are to be conserved at all (see Shepard and others 2005).
Therefore, conservation of native, wild locally-adapted
populations with diverse life history forms will often be
a primary goal. Yet as Meyer (2004) suggested, even in
cases where we cannot conserve an intact evolutionary
legacy, we may still be able to conserve some ecological processes and evolutionary potential. Toward this
end, sustaining wild slightly-hybridized native trout
populations that retain much evolutionary potential to
fill important ecological roles will often be a second
priority, followed by wild populations of non-native trout
that support important ecological and socio-economic
values. Finally, in habitats that cannot sustain wild
trout populations, usually due to lack of reproduction
or recruitment, stocking hatchery fingerlings or adults
may be desired to support valuable fisheries, so long as
these do not threaten ecological or evolutionary values
of other populations (for example, by introgressive
hybridization, or transmitting parasites or pathogens).
This hierarchical approach embodies the principle that
certain values are irreplaceable and should not be forfeited, but that inability to conserve one set of values
does not eliminate the possibility of realizing others.
16
Vulnerability to Invasion
If a native salmonid population of important conservation value is present, the second question to address is
whether it is vulnerable to invasion and displacement
by nonnative salmonids (Table 1). The invasion process
has been divided into three or four basic steps (Vermeij
1996, Kolar and Lodge 2001, Dunham and others 2002a),
which can be summarized as transport, establishment,
spread, and impacts. For simplicity, we discuss transport
and spread together, since they have the same effect of
bringing nonnative salmonids in contact with native
ones. Given this, assessing vulnerability to invasion
requires asking three main questions about the stages
of invasion, shown in Table 2.
Transport and spread—Whether nonnative salmonids are transported to specific locations by humans will
depend on interest in them for recreation or food, and
distances to sources of fish for introduction. Although
most agencies now evaluate fish stocking more carefully than in the past (Rahel 1997, 2000), unauthorized
introductions by the public are burgeoning. For example,
Rahel (2004) calculated that unauthorized introductions
of nonnative fishes in five regions throughout the U.S.
accounted for 90 percent of the new introductions during 1981 to 1999, compared to only 15 to 43 percent
during all previous periods. Disgruntled anglers may
translocate nonnative fishes back to their favorite fishing sites above barriers after they have been removed
using chemicals or electrofishing (U.S. Fish and Wildlife
Service 1998a, Rahel 2004). Established populations
of nonnative salmonids close to native fish populations
(for example, just downstream from accessible barriers,
or a short distance by road) provide a source for such
unauthorized introductions (Thompson and Rahel 1998,
Dunham and others 2004, Rahel 2004, Munro and others 2005).
The spread of nonnative species without the aid of
humans will depend on distance from a source population and the time available for colonization, and on
their tendency for long-distance movement, their rate of
population growth, and the resistance of the corridor to
fish dispersal (With 2002, Hastings and others 2005). If
invasions spread as an advancing front, they will arrive
sooner when source populations are close. However,
nonnative salmonids like brook trout and brown trout
can also move long distances (Young 1994, Gowan and
Fausch 1996, Peterson and Fausch 2003a) and can have
high intrinsic population growth rates (McFadden 1961,
Power 1980, Kennedy and others 2003), both of which
increase invasion rates (Kot and others 1996, Lewis 1997,
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Table 2—A hierarchy of three main questions to assess whether nonnative salmonids will invade and have
negative effects on populations of native salmonids (after Vermeij 1996, Dunham and others
2002a). The three main stages of invasion considered are influenced by additional factors shown
on the right.
Main questions about the invasion process
1. Are invaders likely to be transported or spread to the location that native salmonids inhabit?
Factors influencing invasion process
• Proximity of sources for unauthorized introductions
• Distance to source population and time for invasion
• Tendency for frequent or long distance dispersal
• Population growth rate
• Resistance of the stream corridor to invasion
2. Can the invaders establish a • Environmental resistance - habitat suitability
reproducing population? oTemperature regime
oFlow regime
oStream size
oGradient
oProductivity
• Propagule pressure
• Biotic resistance from native biota
3. Will the invaders displace native
• Biotic interactions and demographic rates may
salmonids by competition, predation, change with abiotic factors (for example,
introgression, or transmission of temperature and flow)
parasites or pathogens, or will they coexist?
• Habitat modification may favor invaders
• Introgressive hybridization may be irreversible
• Parasites and pathogens may be dispersed by invaders
Hastings and others 2005). Therefore, trout invasions
appear to advance through a combination of frequent
short distance movements at the upstream margin
(Rodríguez 2002), combined with less frequent longdistance movements, termed “jump dispersal” (Peterson
and Fausch 2003a). For example, direct measurements
of brook trout movement by tagging and recapture in
Colorado streams revealed that they ranged up to 2 to
3 km within 100-day summer sampling periods, with a
bias toward upstream movement (Riley and others 1992,
Gowan and Fausch 1996). Brown trout movements of
similar distances, and three that reached 23 to 96 km,
were detected by telemetry in Wyoming streams (Young
1994).
Stream segments with high-gradient reaches or
waterfalls may provide resistance to invasion from
downstream, but source populations in headwater lakes
or stream reaches can spread downstream throughout
watersheds. Several syntheses of data from the Rocky
Mountains suggested that brook trout are less successful at establishing populations in streams with gradients steeper than 4 to 7 percent (Fausch 1989, Rieman
and others 1999). However, a mark-recapture study in
Idaho showed that brook trout ascended slopes of 13 to
22 percent, although immigrants in steeper channels
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were larger fish and moved more slowly (Adams and
others 2000). In experimental flumes, adult brook trout
>15 cm could jump over waterfalls up to about 75 cm
high from plunge pools >40 cm deep (Kondratieff and
Myrick 2006), but field results suggest that they may also
be able to ascend higher barriers either by penetrating
interstices at low flow (Thompson and Rahel 1998) or
finding alternate routes at higher flow (Adams and others
2000). In contrast, Adams and others (2001) found that
brook trout dispersed downstream from headwater lakes
where they were stocked, through high-gradient channels
up to 80 percent slope and over an 18-m high waterfall,
potentially invading many reaches inaccessible from
downstream. Likewise, Paul and Post (2001) reported
that brook trout introduced in an Alberta watershed
spread downstream disproportionately from stocking
locations to colonize reaches at lower elevations. There
is also some evidence from northern Montana suggesting
that hybrids of rainbow X cutthroat trout may disperse
more widely than native cutthroat trout (Hitt and others
2003), thereby increasing spatial spread. Rahel (2004)
emphasized that fisheries managers should assume that
any nonnative species that is introduced and becomes
established in a basin will eventually colonize other
suitable and accessible habitats.
17
Establishment—All three main salmonid invaders
have established reproducing populations in most U.S.
states where they are not native (Rahel 2000), although
they have not invaded all waters to which they have
access (Fausch and others 2001, Adams and others
2002, Dunham and others 2002a). Establishment of
a nonnative salmonid population will depend on both
environmental resistance (habitat suitability) and “biotic
resistance” by native species, as well as chance events
(Moyle and Light 1996). Habitat suitability for native
and nonnative salmonids in streams is often governed
by temperature regime (Meisner 1990, Rahel and Nibbelink 1999, Harig and Fausch 2002, Dunham and others
2003a), flow regime (Jowett 1990, Jager and others 1999,
Strange and Foin 1999, Fausch and others 2001), stream
size (Rahel and Nibbelink 1999, Rieman and others
1999), habitat factors correlated with gradient (Fausch
1989, Adams 1999), and stream productivity (Kwak and
Waters 1997). Because many of these factors influence
fish at thresholds, measurements across a wide range of
each variable often show that establishment is relatively
predictable from simple physiological tolerances (see
Box 5). In other cases, simultaneous consideration of
a host of variables may be required (Moyle and Light
1996, Marchetti and others 2004).
A main predictor of success of deliberate introductions or translocations of nonnative species is “propagule
pressure” (Kolar and Lodge 2001, Hilderbrand 2002,
Marchetti and others 2004), which refers to the number of
releases and the number of individuals per release. These
data indicate that repeated stocking of nonnatives can
eventually increase likelihood of establishment. In part,
this may be due to rare events, such as years with unusual
weather, which favor nonnative species and allow them
to establish “beach head” populations and subsequently
invade (Moller 1996, Moyle and Light 1996).
Intact assemblages of native fishes may also provide
biotic resistance that prevents nonnative salmonids from
invading, especially when the assemblages include more
species. Current theory in invasion biology holds that
assemblages with more species provide fewer “niche
opportunities” for invaders (Shea and Chesson 2002,
Chase and Liebold 2003), defined as opportunities for
invading species to increase from low densities due to
abundant unused resources or a lack of natural enemies.
This theory may explain why brook trout in western
North America are least successful at invading Pacific
Northwest coastal watersheds that support intact runs
of salmon and anadromous trout (Oncorhynchus spp.)
with many diverse life history types that use a wide
variety of food and space resources.
18
Box 5. Examples of environmental resistance that
limits invading salmonids.
Several examples indicate that abiotic factors like
low temperature and flooding regimes can limit
establishment of nonnative salmonids, but each may
interact with other factors.
Low temperature: Low temperature limits salmonid
recruitment by delaying egg development and hatching which causes higher egg mortality (Stonecypher and others 1994), and by reducing fry growth
and lipid storage which causes higher overwinter
mortality due to starvation (Biro and others 2004,
Coleman and Fausch 2006). Low temperature also
limits growth of adults, which reduces fecundity
and recruitment.
•Rahel and Nibbelink (1999) found that upstream
limits of brown trout in a Wyoming watershed could
be explained by a combination of cold temperature
and small stream size.
•Adams (1999) reported that reduced fecundity at
low temperatures could account for the upstream
limit of brook trout in a Montana stream.
•Clark and others (2001) inferred that lower
­fecundity due to lower temperatures limited invading rainbow trout in the northern Appalachian
Mountains.
Flooding regime: Flooding during incubation or
emergence of fry from gravel redds may limit establishment of nonnative salmonids.
•Nehring and Anderson (1993) and Latterell and
others (1998) found that recruitment of nonnative
brook, brown, and rainbow trout in sets of Colorado
rivers was lowest in years when snowmelt runoff
flows were highest, probably because fry emerge
just before or during these flow peaks and are
displaced.
•Fausch and others (2001) reported that nonnative
rainbow trout were least likely to establish reproducing populations in several regions worldwide
where flooding occurred during fry emergence in
early summer, and most likely to establish in regions
where flow regimes most closely matched those in
their native range along the Pacific coast of North
America.
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Displacement, coexistence, and other effects—In
many cases, after nonnative salmonids are transported
or spread to a stream and establish a reproducing population, they reduce or displace native salmonids (Waters
1983, Behnke 1992, Peterson and others 2004), but in
other cases the invasion may stall or the two may coexist
(Shepard 2004, Rieman and others 2006). Adams and
others (2002) reported that upstream limits of brook
trout distribution and recruitment advanced over a 25year period in fewer than half the streams resampled
in an undisturbed Idaho watershed. Whether displacement or coexistence occurs may depend on abiotic
factors like temperature that affect biotic interactions
among individuals at the local scale, or on temperature
and flow regimes that affect demographic responses
at larger scales. For example, at the local scale, DeStaso and Rahel (1994) reported that brook trout were
behaviorally dominant over Colorado River cutthroat
trout at warmer temperatures in an artificial stream,
whereas neither was dominant at colder temperatures
(but see Novinger 2000). However, it is more likely that
the effects of temperature and other abiotic factors will
play out at riverscape scales by changing demographic
rates. For example, low temperature may reduce growth
or fecundity of nonnative salmonids and thereby limit
their latitudinal distribution (Clark and others 2001) or
upstream invasion (Adams 1999, Weigel and others 2003,
Benjamin 2006). Alternatively, changes in environmental
conditions that favor productivity of non-native species
might allow them to invade new habitats and displace
native species. Moller and Van Kirk (2003) reported
that nonnative rainbow trout have been more successful
at invading and displacing native Yellowstone cutthroat
trout (O. c. bouvieri) where flow regimes were modified to favor rainbow trout over cutthroat trout. Shepard
(2004) suggested that brook trout were more likely to
displace westslope cutthroat trout where habitat had been
degraded by land management activities. Likewise, Olsen
and Belk (2005) suggested that native fishes were less
likely to coexist with non-native predators such as brown
trout if habitat had been degraded so that off-channel
refugia were lost. These examples also suggest that judicious management of flow and temperature regimes, and
physical habitat structure, to achieve more natural states
might favor native species (Poff and others 1997, Fausch
and others 2001), and allow some level of coexistence
with nonnative species. This idea would benefit from
careful study using an adaptive management approach
at watershed scales (see Walters 1986).
Two other possible effects of invaders are hybridization
and introduction of parasites or pathogens. Although
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invasions of nonnative salmonids can sometimes be
eradicated (Moore and others 1983, Knapp and Matthews 1998), especially if caught in the early stages
(Simberloff 2003), neither hybridization nor introduction of parasites or pathogens to a native salmonid
population is usually reversible. Only in cases where
hybrids have substantially lower survival or fertility, as
perhaps for brook trout hybridizing with bull trout (see
above), could hybridization be reversed. Introductions of
parasites like Myxobolus cerebralis are also apparently
irreversible once established, and may be transmitted
farther by movements of invading salmonids upstream
or downstream from infected source populations (see
above; Nehring and Thompson 2003, Nehring 2004).
However, whether conditions are favorable to persistence and expansion of pathogens may depend on local
environmental conditions, and even past watershed
management (Anlauf 2005).
Isolation and Population Persistence
If a native salmonid population is vulnerable to a
nonnative salmonid invading from downstream, one
management alternative is to construct a barrier to fish
movement to prevent the invasion. However, then the
influence of isolation caused by the barrier on persistence of the native salmonid population becomes an
important question (Table 1). Isolation or fragmentation
of habitat has several important effects on native salmonids (Table 3; Rieman and Dunham 2000, Dunham
and others 2003b), including declines in population and
habitat size. Small populations can be more vulnerable
to extinction because of a loss of genetic variability,
random changes in demographic processes and rates
(demographic stochasticity), and environmental fluctuations (environmental stochasticity), collectively known
as “small population” phenomena (Caughley 1994).
Large-scale perturbations or “catastrophes” that severely
reduce populations and habitats can be important for
both small and large populations, particularly if they are
confined to a set of spatially restricted habitats that could
be affected by the same disturbance, such as hurricanes,
earthquakes, fires, floods, droughts, and temperature
extremes (Propst and others 1992, Brown and others
2001). Moreover, disturbances that pose little threat to
larger, interconnected populations may become a threat
when populations are highly fragmented (Dunham and
others 2003b).
Isolation also eliminates migratory life histories,
which confer resilience to local populations. Populations
with migratory individuals are spread across many different habitats and may therefore be less vulnerable to
19
Table 3—Basic threats to the persistence of single, isolated salmonid populations, and
guidelines for addressing each threat (see Rieman and McIntyre 1993, Allendorf
and others 1997, McElhany and others 2000 for more detailed criteria). Criteria
for multiple populations are given elsewhere (see Box 6).
Threat
Guidelines
Loss of genetic variability
Maintain enough adults to minimize loss of genetic
­variability through drift and inbreeding.
Loss of resilience
Maintain or restore the condition of local habitats or
reduce sources of mortality to compensate for lost
­fecundity and potential productivity due to loss of
­migratory adults.
Demographic stochasticity
Maintain enough adults and year classes to minimize
the probability that chance demographic events drive
local populations extinct.
Environmental stochasticity
Maintain a large enough population size and wide
­distribution within the isolated habitat to persist in the
face of common environmental fluctuations.
Ensure that human influences do not result in habitat
degradation or substantial changes to environmental
regimes that reduce survival.
Catastrophes
Maintain a large enough habitat and population size, and
wide distribution, to ensure resilience in the face of rare,
high-magnitude environmental changes.
Ensure that human influences do not alter the frequency
and magnitude of natural catastrophes in a way that
threatens remnant populations, or create new catastrophes
(for example, toxic chemical spills, introductions of invasive
species) that pose threats.
disturbances influencing a single habitat. For example,
short-term local extirpations of cutthroat trout by drought
(Dunham and others 1997) and bull trout by wildfire
(Rieman and others 1997a) were likely reversed by returns of migratory fish that were in habitats outside of
the systems during the disturbance. Isolation or other
influences that lead to loss of migratory life histories
restrict a population spatially and thus increase the vulnerability to local disturbances. Isolation also reduces
the opportunity for exchanges of individuals among
local habitats, which may reduce the demographic stability of some populations (Hilderbrand 2003). Finally,
migratory fish are typically larger and more fecund,
which may provide important demographic support to
existing populations, and possibly resistance to invading
salmonids (Dunham and others 2002a).
Threats posed by isolation cannot be considered alone.
Synergistic feedbacks among threats can lead to the
20
demise of local populations (for example, an “extinction vortex,” Gilpin and Soulé 1986). Many factors can
modify population responses of salmonid fishes to isolation and strongly influence local persistence, including
angling, habitat degradation, and habitat and life history
characteristics. The negative effects of angling on native
trout populations can be substantial, even if regulations
restrict harvest (Post and others 2003), particularly for
large migratory fish that are highly desirable to anglers.
Habitat degradation often leads directly to habitat fragmentation in stream networks, and also reduces survival
and population size making populations less resilient and
more vulnerable to other threats. Habitat degradation can
influence salmonids at a variety of spatial scales. Local
factors that are strongly altered by land management,
such as temperature, sediment, stream discharge, and
instream cover, are commonly cited as degrading habitat
(Bjornn and Reiser 1991). At broader scales, the status of
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salmonid populations has been consistently associated
with the distribution of land management activities, including road density and management intensity (Thurow
and others 1997, Baxter and others 1999). Habitat and
life history characteristics can also influence the probability of salmonid population persistence in the face of
isolation. For example, Koizumi and Maekawa (2004)
observed that Dolly Varden charr were more likely to
persist in smaller isolated habitats if they were strongly
influenced by groundwater, which provided more stable
and favorable conditions over time. Rieman and Dunham
(2000) argued that cutthroat trout were more resistant to
local extinction in small habitat patches than were bull
trout. These examples illustrate that other environmental
conditions and species characteristics may mitigate or
exacerbate the effects of isolation.
Considering Priorities among Multiple
Populations
To this point, our framework has addressed the tradeoff
of invasion and isolation for single populations. Fisheries
managers and conservation biologists have no lack of
problems to consider at this scale, but often move from
one project (that is, the implementation or removal of
a barrier) to the next as the need or opportunity arises.
Typically, this is done while struggling with limited
financial, capital, and human resources. However, to
be most effective it is important to avoid incremental
or “ad hoc” solutions and carefully prioritize use of
limited resources (Pressey 1994, Allendorf and others
1997, Groves 2003). This can be facilitated through a
systematic assessment of risks (Francis and Shotton
1997). Characterizing risk implies understanding: 1)
the probability of an event (for example, a nonnative
salmonid invasion), 2) the effects of the event if it actually occurs (for example, probability of ecological displacement or genetic introgression following invasion),
and 3) the value that could be lost. Thus, a “threat” is
a factor that could negatively impact a population, and
a “risk” is the probability that a given factor will have
an impact on something we value.
A simple example can illustrate these three components
of risk for native cutthroat trout. First, the probability of
invasion may be higher for cutthroat trout populations
closer to sources of invasion by brook trout, increasing
the first component of risk. Second, the effects of invasion may be greater for some cutthroat trout populations
because they are less resistant to invasion, either due to
their weaker population status or because they live in
habitats that favor brook trout, increasing the second
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component of risk. Third, the conservation value of some
cutthroat trout populations may be higher than others
because they represent relatively rare pure genomes or
life history types, increasing the third component of risk.
In contrast, populations farther from invasion sources,
more resistant to invasion, or of lower conservation
value would be at lower risk. Therefore, these would be
lower priorities as candidates for constructing barriers
to prevent invasions. These are simple examples, and
tradeoffs can become more complicated as multiple factors and populations, spatial structure, and the general
principles of conservation planning are considered.
The first step in planning is to identify the fundamental units of conservation and determine how they are
arranged across spatial scales. Native salmonids often
occur as collections of local populations distributed
across habitats that are nested within larger subbasins
(Rieman and Dunham 2000). This creates a hierarchy of
spatial structure at scales ranging from local “patches”
of suitable habitats (Dunham and others 2002b) to patch
networks, and even larger aggregations or regional
networks (Table 4). Moreover, some habitats may be
potentially occupied but currently vacant. We emphasize
that “scale” as used here should be based on how fish
populations are organized across the landscape, and not
on a fixed spatial scale or stream network structure. In
many cases, the natural structure of local populations
has been modified by human activities, such as barriers
that fragment habitats. In other cases for rare taxa, only
one local habitat exists, such as for the Paiute cutthroat
trout (O. c. seleniris). Whatever the case, it is important
to carefully consider the biology of the target species and
how it interacts with the larger landscape, to define local
populations, habitat patches, patch networks, or other
units for use in conservation planning (Dunham and
others 2002b, Fausch and others 2002, Wiens 2002).
The second step is to establish priorities for management of the potential conservation units. Ultimately,
managers will need to decide on the number, distribution, and characteristics of populations they will attempt
to conserve and the actions required (that is, installing
or removing barriers). The conservation literature is replete with proposals outlining goals for such efforts, but
the most common include representation, redundancy,
persistence, and feasibility (Scott and Csuti 1997, Groves
2003). Moreover, a formal decision process must also
acknowledge that setting priorities is inherently uncertain
(Ludwig and others 1993, McElhany and others 2000). In
practice, these goals do not always translate neatly into
specific guidelines (see Box 6), but we discuss each separately here to emphasize their individual importance.
21
Table 4—Potential relationships between conceptual scales at which natural populations of salmonid
fishes are structured (Dunham and others 2002b) and units of conservation as defined in
practice (following the Draft Bull Trout Recovery Plan; U.S. Fish and Wildlife Service 2002).
Conceptual Scale
Draft Bull Trout
Recovery Plan
Patch
Local Population
A patch may be occupied or not at present,
whereas local populations are based on
occupancy. Occupied patches and local
populations are characterized by frequent
(daily to seasonal) interactions among individuals.
Patch network
Core Area
or metapopulation
Local aggregation of patches or local
populations characterized by less frequent
(annual to decadal) interactions.
Subbasin
Recovery Unit
Naturally discrete aggregations of patch
networks or core areas within larger drainage
basins that potentially interact over long time
scales (100s to 1000s of years).
Region
Major biogeographic units that characterize
distinct evolutionary lineages.
Distinct Population Segment
Representation refers to how well local populations
and habitats reflect the full suite of ecological and evolutionary conditions that are expressed within a given
area (Lawler and others 2003, Ruckelshaus and others
2003). In conservation planning, it is nearly always the
case that only some of the “pieces” can be protected or
restored. Accordingly, it is important to identify and
protect those that provide the most complete representation of the original population characteristics (for
example, allele frequencies, life history types, species
occupancy or composition) and habitat conditions (for
example, spatial distribution, successional stage), as well
as the conservation values defined above.
Redundancy is a key component of conservation planning because no local population or habitat is immune to
extinction or catastrophic events (Moyle and Sato 1991).
Therefore, it is prudent to “not put all your eggs in one
basket,” and instead to conserve multiple examples of all
the representative pieces. Although many have struggled
to identify the minimum number and distribution of
local populations necessary to insure persistence of a
metapopulation (see Rieman and McIntyre 1993) that
number will always depend on the local environments
and events that are difficult to predict. In general, more
populations will be necessary as individual populations
become more vulnerable (for example, smaller), the
potential for catastrophic disturbances increases, or the
potential for simultaneous disturbance across sets of
22
Description
local populations increases (Hanski and Gilpin 1997).
Replicate populations are especially important when
all are isolated above barriers, to allow reintroduction
if local extinctions occur but habitats remain suitable
(Lubow 1996).
Persistence is a function of the resilience of individual
populations, as well as the larger network of populations (Table 4). Factors contributing to the persistence
of individual populations in the face of isolation and
invasion were considered above, and have been addressed in detail for salmonids (Rieman and McIntyre
1993, Allendorf and others 1997, McElhany and others
2000). Populations that appear unlikely to persist may
be low priorities for conservation actions. Persistence
of multiple populations must usually be considered
at several scales. For example, the dynamics of local
habitats and populations may create a metapopulation,
which may be nested within a larger collection of core
areas that characterize a recovery unit, and ultimately
a distinct population segment (DPS) or evolutionarily
significant unit (ESU; Table 4). Whereas the spatial
structure and temporal dynamics may be difficult to
quantify precisely (Rieman and Dunham 2000, Fausch
and others 2002), it is important to consider these factors because persistence of any single population may
depend strongly on influences from adjacent populations
or habitats. We strongly encourage readers to consult
key references (Rieman and McIntyre 1993, Allendorf
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Box 6. Guidelines for viability of multiple salmonid populations (partially excerpted and modified from
McElhany and others 2000; see also Rieman and others 1993, Allendorf and others 1997).
• Some populations should exceed minimal guidelines for long-term viability. Larger and more productive
(“resilient”) populations may be able to recover from a catastrophic event that would cause the extinction of smaller
populations. Furthermore, such populations may serve as important sources of dispersing fish to less productive
habitats. A collection of local populations that contains some populations with high abundance and productivity is
less likely to go extinct in response to a single catastrophic event that affects all populations. It should be recognized
that populations or habitats serving as sources and sinks may exchange roles over time.
• Populations that display diverse life-histories and phenotypes should be maintained. When local populations
each have a fair degree of life history diversity (or other phenotypic diversity), a collection of local populations is
less likely to go extinct as a result of correlated environmental catastrophes or changes in environmental conditions
that occur too rapidly for an evolutionary response. In addition—assuming phenotypic diversity is caused at least
in part by genetic diversity—maintaining diversity allows natural evolutionary processes to operate. Note that to
protect genetic diversity, it may be necessary to maintain several populations with the same phenotype.
• Habitat patches should not be destroyed faster than they are naturally created. Salmonid habitat is dynamic;
suitable habitat is continually being created and destroyed by natural processes. Human activities should not decrease
either the total area of habitat OR the number of habitat patches.
• Maintain some habitat patches that appear to be suitable or marginally suitable, but currently contain no
fish. In the dynamics of natural populations, there may be time lags between the appearance of empty but suitable
habitat (by whatever process) and the colonization of that habitat. If human activity is allowed to render habitat
unsuitable when no fish are present, the population as a whole may not be sustainable over the long term.
• Natural rates of straying among subpopulations should not be substantially increased or decreased by
human actions. This guideline means that habitat patches should be close enough together to allow dispersal and
expansion of the population into under-used patches during times when productivity is high. Also, dispersal should
not be much greater than natural levels because increased dispersal may reduce a population’s viability if fish wander
into unsuitable habitat or interbreed with genetically unrelated fish.
• Some populations should be geographically widespread. Spatially correlated environmental catastrophes are
less likely to drive a widespread collection of local populations to global extinction.
• Some populations should be geographically close to each other. On long temporal scales, having populations
geographically close to one another facilitates connectivity among existing populations. Thus, a viable collection
of local populations requires both widespread AND spatially close populations.
• Populations should not all share common catastrophic risks. A collection of local populations that do not
share common catastrophic risks is less likely to be driven to extinction by correlated environmental catastrophes.
Maintaining geographically widespread populations is one way to reduce risk associated with correlated catastrophes, but factors unrelated to spatial proximity may be important. For example, geographically distant populations
with similar vulnerability to broad-scale climate influences may share common catastrophic risks.
• Evaluations of status should take into account uncertainty about among-population processes. Our understanding of spatial and temporal process and interactions among local populations is very limited. Because most
collections of local populations of salmonid fishes are believed to have been historically self-sustaining, the historical
number and distribution of populations serves as a useful “default” goal, unless better evidence is available.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
23
and others 1997, McElhany and others 2000) to develop
individual applications. Although throughout this monograph we address persistence of populations, we realize
that the broader goal of viability may more often be the
conservation target of interest (see Box 7).
Feasibility means that conservation actions should be
cost effective, sustainable, and socially and environmentally acceptable. If all else is equal, projects leading to
the most efficient implementation or removal of barriers
should be those that will produce the greatest benefit (for
example, reduction in risk) for the least cost. Cost must
be considered not only for the immediate project, but
also for long-term maintenance. Constructing barriers or
replacing impassible road culverts with ones that allow
fish passage is costly, so only a few can be implemented
each year. Efforts to control or eradicate nonnative
salmonids have proven expensive and are not always
successful (Dunham and others 2002a), so thoughtful
evaluation of the cost-effectiveness of these efforts will
be important. Constructing or removing ­ barriers can
Box 7. Persistence vs. viability.
Often the terms “persistence” and “viability” are
used interchangeably, but here we recognize viability
as a larger conservation objective. The basic impetus
for conservation planning is not simply to guarantee
persistence of a species, but to ensure that natural
ecological and evolutionary processes are allowed
to continue and perhaps change through time. For
a single species, this broader view of maintaining
process, not just persistence, is referred to here as
“viability” (see McElhany and others 2000). In the
prioritization process, this may equate to conservation of both evolutionary and ecological values
simultaneously. For example, evolutionary values
can be associated with genetically pure but isolated
populations that may persist for some time, but if
those populations cannot evolve and adapt with
changing environments they may not be viable in
the long term. Populations that are likely to persist
and remain viable represent a higher overall value
and logically a higher priority in any assessment of
risk. In short, persistence is generally viewed as a
necessary but not sufficient objective for attaining
full conservation of a species.
24
have other environmental effects as well. Barriers to
non-native fish are also barriers to other aquatic organisms that may need to move to persist and maintain
productive aquatic and riparian communities. Barriers
also disrupt flows of water and materials (for example,
sediment) that shape upstream and downstream habitats
or affect floodplains and human infrastructure.
Finally, it is clear that conservation planning decisions
are inherently uncertain. For example, our understanding
is very limited about where brook trout will invade, how
long isolated populations of native fishes will persist,
and how these differ among different environments and
species or subspecies. This creates genuine disagreement among biologists who have experience in different
regions. Although these limitations can be frustrating,
we will never have all the answers and decisions must
be made, so confronting uncertainty must become part
of the management process (Ludwig and others 1993,
Francis and Shotton 1997, McElhany and others 2000).
Ludwig and others (1993) suggested that managers should
favor decisions that are robust to uncertainty (that is,
the outcome is likely to be favorable regardless of our
knowledge). When this is impossible, it is important to
hedge, probe the system to learn, and favor reversible
actions. For example, hedging could include maintaining redundant populations and constructing barriers in
some systems but not others to learn more about what
happens, through careful monitoring. Some barriers
may also be viewed as short-term, reversible actions to
prevent impending invasions, while learning more about
the threats of isolation. We are uncertain when the effects
of barriers become irreversible because, for example,
migratory forms are lost. Recent work suggests that the
evolution or re-expression of migratory life histories
and colonization of newly accessible habitats can occur
quickly (Hendry and Stearns 2004, Quinn 2005). However, until research or management experiments clearly
demonstrate that the effects of isolation can be reversed,
any intentional isolation must be viewed as a calculated
risk, balanced against the threat of invasion.
Section VI: Two Examples of
Invasion and Isolation Tradeoffs____
We use the Coeur d’Alene River basin in northern
Idaho and the Little Snake River basin in southern
Wyoming and northern Colorado (Table 5) as examples
of relatively complex and simple systems, respectively,
to show how our conceptual framework can be used to
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Table 5—Aspects of the physical characteristics, trout populations, and management issues in the
Coeur d’Alene River basin, Idaho, and Little Snake River basin, Colorado-Wyoming.
Characteristic
Coeur d’Alene River
Little Snake River
Basin area (ha)
232,112
288,344
Elevation range (m)
600 to 2,1001,900 to 3,300
Period of peak precipitation (form)
November to January
(snow and rain)
March to April
(snow)
Primary land management activities
mining, logging
grazing, water development
Cutthroat trout subspecies
westslope cutthroat trout
Colorado River cutthroat trout
Life histories present
resident, fluvial, adfluvial
resident
Cutthroat trout distribution (km)1,063160 to 190
Other sympatric native salmonids
bull trout1
mountain whitefish
mountain whitefish1
Sympatric nonnative salmonids
brook trout, rainbow trout
none2
Brook trout distribution (km)92
unknown but widespread
and much greater than
cutthroat trout
Ranking of cutthroat trout
population values
socio-economic >
ecological > evolutionary
evolutionary > ecological >
socio-economic
Management emphases
maintain fishery for
fluvial and adfluvial fish
create and enlarge remnant
cutthroat trout populations
Management tactics
restrict harvest, improve
habitat, remove barriers,
install barriers, stock
sterile rainbow trout
remove brook trout, install
barriers, translocate
cutthroat trout
1
Neither cutthroat trout subspecies currently coexists with these salmonids.
Colorado River cutthroat trout largely persist only above barriers preventing invasions of nonnative trout, but individuals occasionally drift downstream over barriers to form sympatric assemblages.
2
consider the tradeoffs in managing invasions and isolation. There are many factors that affect conservation of
native salmonids, including the history and extent of
land management, location of natural and anthropogenic
barriers to fish movement, introduction and invasion of
nonnative salmonids, and goals of fisheries managers.
These issues vary markedly across the western U.S., and
it is beyond our scope to explore them in great detail
here. Our goal is to use the framework outlined above
to consider the opportunities and priorities that emerge
in each case. However, we caution that these are not
intended to be specific proposals for management.
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Coeur D’Alene River
Background—The upper Coeur d’Alene River basin
in northern Idaho (Figure 6) has two major branches,
the North Fork and Little North Fork Coeur d’Alene
rivers, each with numerous 2nd- to 4th-order tributaries.
The basin has relatively low topographic relief (approximately 600 m), but is influenced by a mixture of climatic
processes. The hydrograph is predominantly driven by
melting snow, but the Pacific maritime influence and
the relatively low elevation of much of the basin result
in some winter rain. Rain-on-snow floods during winter
25
Figure 6—Predicted distribution of spawning and rearing habitat for westslope cutthroat trout (bold stream
segments), and road culverts believed to be complete barriers to all fish passage (dots). Spawning and rearing
habitat was predicted from associations between young-of-the-year cutthroat trout and stream size (Dunnigan
1997, Abbott 2000), gradient, and confinement (Moore and Gregory 1988, Lentz 1998), and from direct field
observations of spawning adults (J. Dupont, Idaho Fish and Game, personal communication). Barriers were
identified by the Panhandle National Forests stream crossing inventory. The dotted line encompasses the area
believed to support cutthroat trout populations that have not been hybridized by rainbow trout. Some segments
above barriers outside this area may also retain pure cutthroat trout. One headwater tributary in the northeast
portion of the basin has hybridized trout and so was excluded from this area.
storms are more common than in much of the range of
westslope cutthroat trout. There are no headwater lakes
in the basin that serve as sources for downstream invasion or modify the hydrologic and thermal regimes of
downstream habitats.
Disruptive management and development have been
extensive. Hard rock mining, dredging, log drives down
streams, clearcut logging, extensive road building, and
overfishing have degraded habitats and depressed trout
populations, especially in the lower half of the basin.
In addition, road crossings have created barriers to
fish movements in many tributaries. Native salmonids
included westslope cutthroat trout, mountain whitefish
(Prosopium williamsoni), and bull trout, although bull
26
trout are now presumed to be extirpated. Because of its
unusual geomorphic and climatic characteristics (for
example, heterogeneous geology and frequent winter
flooding), the Coeur d’Alene River basin could represent
a unique evolutionary template for westslope cutthroat
trout (for example, Allendorf and others 1997). Resident
and migratory trout occur throughout the basin, including fluvial and adfluvial fish that reach 350 to 450 mm
(14 to 18 inches). Brook trout and rainbow trout were
introduced between 1900 and 1950, and both have become established in parts of the system. Hybrid swarms
of cutthroat trout X rainbow trout are common in the
lower tributaries, but hybrids declined or were absent
entirely in samples upstream.
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Restrictive angling regulations (that is, size limits, gear
restrictions, and catch-and-release) were first imposed on
portions of the basin in the 1970s. Hatchery rainbow trout
have been stocked in the main-stem rivers for decades,
but less so in recent years. Now, only sterile fish are
planted to reduce the risk of further hybridization. Habitat
management has focused on road obliteration, channel
reconstruction, and barrier removal. The trout populations
have responded and the upper river has recently become
a renowned fly fishing destination for wild trout.
Conservation values—Evolutionary, ecological, and
socio-economic conservation values are all considered
important in the Coeur d’Alene River basin, but the last
two are dominant management objectives. A productive, diverse, and resilient population of trout with a
large migratory component is an emphasis of natural
resource managers. The sport fishery for large cutthroat
trout in the main-stem rivers plays a significant role
in the economy of local communities, particularly as
traditional extractive industries (for example, timber
harvest and mining) are replaced by tourism.
Conservation of evolutionary value (which might be
compromised by hybridization) has been a secondary
concern, primarily because hybridized trout that may
be morphologically indistinct from genetically pure cutthroat trout are widespread and have been considered
part of the formally recognized taxonomic group (U.S.
Fish and Wildlife Service 2003, Campton and Kaeding
2005). There is still controversy on this issue (Allendorf
and others 2004, 2005). Little work has explored the different ecological roles of cutthroat trout, rainbow trout,
or their hybrids, but abundant populations of the latter
two can become established in habitats once occupied
by cutthroat trout. A general assumption seems to be
that rainbow trout or their hybrids may persist, evolve
ecological roles, and support fisheries that are surrogates
for native cutthroat trout, with little loss of the associated ecological functions and values (see Quist and
Hubert 2004). However, although native and non-native
trout are likely to have similar roles at some levels of
ecological organization (McGrath 2004), in other cases
non-native trout have very different effects (Baxter and
others 2004). Little research has been done to clarify
the ultimate ecological implications of introgression or
displacement of cutthroat trout by other salmonids, so
some undefined risk remains.
To illustrate the fundamental units of conservation and
the distribution of local populations needed to maintain
all conservation values, we mapped the potential distribution of spawning and rearing habitat for cutthroat trout.
We assumed that natal habitats define local populations
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(Table 4), and represent the critical habitats vulnerable
to invasion and isolation. We used simple models based
on landscape and channel characteristics associated
with the distribution of juvenile fish (Dunnigan 1997,
Abbott 2000) and observations of spawning areas by
local fisheries biologists. We concluded that stream segments potentially important to conservation of ecological
and socio-economic values occur in tributary systems
throughout the basin (Figure 6).
Segments potentially important for maintaining evolutionary value are more restricted. Extensive hybridization
is common in the lower tributaries, but pure populations
are found in the upper portions of both the Little North
Fork and North Fork subbasins (B. Rieman, unpublished
data). Therefore, it seems likely that genetically pure
populations of westslope cutthroat trout important for
conserving evolutionary values also support socioeconomic and ecological values in the upper part of
the basin. Road crossings have also created barriers to
movements of trout in many tributaries (Figure 6), and
these may have temporarily prevented the invasion of
rainbow trout and the extent of hybridization. Genetically
pure populations of cutthroat trout may persist above
some barriers, but surveys are needed to confirm this.
Vulnerability to invasion—Brook trout are well
established in some tributaries of the Little North Fork
and the lower North Fork Coeur d’Alene rivers. To
consider habitats most vulnerable to future invasion
and displacement of westslope cutthroat trout, we used
an approach similar to that described above to map
potential habitat for brook trout (Figure 7). Our predictions suggest that brook trout could ultimately occupy
and potentially displace cutthroat trout from portions of
many tributaries throughout the basin. If the invasion
proceeded to the ultimate limits it could leave habitat for
cutthroat trout highly fragmented. A complete invasion
would threaten all three conservation values associated
with the existing trout populations, although brook trout
might still fill some of the ecological roles played by
salmonids in these streams (see McGrath 2004). The
most vulnerable habitats are tributaries adjacent to, or
habitats upstream of, current populations of brook trout
where propagule pressure should be highest. However,
brook trout remain patchily distributed, even though
they have been established in some tributaries for at
least 60 years (Maclay 1940). Brook trout and cutthroat
trout also appear to coexist in some streams. Whether
the brook trout invasion is stalled or advancing very
slowly is unknown, so it also is not clear that invasion
and displacement is imminent even in sites close to
potential source populations.
27
Figure 7—Known (red stream segments) and potential (yellow segments) distribution of brook trout in the
upper Coeur d’Alene River basin. Potential brook trout habitat was predicted based on associations between
brook trout distributions and stream size and channel gradient (Fausch 1989, Rieman and others 1999, Rich
and others 2003, Petty and others 2005). However, brook trout could invade farther than indicated here
because the models are simple approximations and some fish may disperse beyond their primary habitats. The
brook trout distribution in the Coeur d’Alene River basin does not appear to have advanced in recent decades,
so other factors like temperature or winter flooding may limit the distribution to less than predicted here.
The threat of invasion by rainbow trout and their hybrids appears more immediate than invasion by brook
trout. Rainbow trout or their hybrids could ultimately
invade throughout the system even without further
stocking of fertile rainbow trout because hybridization
may facilitate dispersal (Hitt and others 2003, Allendorf
and others 2004). Thus, conservation of evolutionary
values associated with pure populations of cutthroat
trout in the headwaters or any accessible tributaries
throughout the basin is threatened. The distribution of
rainbow trout genes may be constrained by elevation or
thermal gradients (Weigel and others 2003) so the risks
of hybridization are probably greater for populations
closer to the advancing front. In contrast, conservation of ecological and socio-economic values is of less
concern because hybridized trout may retain many of
these values.
28
Persistence with isolation—Westslope cutthroat trout
have persisted in isolated stream segments throughout
the region, but little information is available to directly
estimate the probability of persistence as a function of
isolation time, habitat condition, or network size. Without
data for this taxon we can only assume westslope cutthroat trout face threats similar to those outlined in the
preceding discussions for other salmonids. Consequently,
we anticipate that stream networks with less than about
10 km of connected suitable habitat will be at increased
risk of local extinction from stochastic, demographic,
and genetic processes. The risks will increase as network
size declines, habitat is degraded, or the potential for
catastrophic disturbance increases. Populations that are
isolated will also lose any demographic resilience associated with migratory life histories unless the isolated
network is large enough to support more productive
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main-stem habitats that serve as foraging areas supporting faster growing individuals. Moreover, the basin
is vulnerable to winter rain-on-snow floods associated
with slow-moving, moist winter storms, which can create
catastrophic disturbances especially in the aftermath of
large fires. These events could drive small populations
extinct, so any strategy depending on intentional isolation
should consider spatial replication across broad scales
to reduce the potential for simultaneous extinctions.
Potential priorities and actions—Priorities for management in this basin might include both removing and
constructing barriers as well as improving habitats to
enhance the resilience of local populations. Because there
is still opportunity to provide some representation and
diversity of all primary conservation values, a mix of
strategies could be important. In general, constructing
barriers could be used to conserve evolutionary values
in remaining genetically pure populations, whereas removing barriers could be used to conserve and expand
the potential distribution of large migratory trout that
represent important ecological and socio-economic
values.
There may be an unusual (albeit costly) opportunity to
conserve all three conservation values simultaneously.
The fishery and resilience of the remaining cutthroat
trout populations depend strongly on migratory life
histories and the connection of tributary and main-stem
habitats. However, these connections could allow both
brook trout and rainbow trout to expand throughout
the system. The brook trout invasion appears most
imminent in the Little North Fork subbasin, whereas
the risks are probably lower in the North Fork where
the distance from potential sources is much greater.
In contrast, the rainbow trout invasion may extend
throughout the system. Large barriers might be used to
conserve evolutionary as well as ecological and socioeconomic values if they could isolate tributary-river
networks where pure populations and migratory life
histories still persist (Figure 8). The best opportunity
for this appears to be in the headwaters of both river
subbasins where pure populations and multiple tributary
networks are linked to the main-stem rivers. Intentional
barriers could be installed first in the headwaters of the
Little North Fork Coeur d’Alene River because brook
trout and rainbow X cutthroat trout hybrids are present
immediately downstream (Figures 6, 7, and 9). If this
proved successful and invasions continued to advance
in the North Fork, large barriers could be installed with
the benefit of the experience gained in the first project.
Although large barriers like low-head main-stem dams
and diversions are possible, they may not be feasible.
Large barriers would be expensive, would disrupt important ranging movements of large fish and potentially
other species in the upper rivers (J. Dupont, Idaho Fish
Figure 8—Pure populations of large fluvial westslope cutthroat trout (Edward Lider, Panhandle National Forest, image)
might be conserved in the upper North Fork Coeur d’Alene River basin by using barriers to prevent upstream invasion
and hybridization by nonnative rainbow trout. If this were successful, the native cutthroat trout would support al three
types of conservation values.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
29
and Game Department, Coeur d’Alene, unpublished
data), and could have negative effects on hydrologic and
geomorphic processes. Barrier management of this scale
would require significant commitment of resources and
environmental compromises.
More modest alternatives might include strategic use of
barriers to conserve representative genetic diversity and
evolutionary values in smaller stream networks throughout the basin. Existing impassable culverts have isolated
several relatively large (5 to 10 km) habitat networks
across the system (Figure 9). If remnant populations
above these barriers have not been hybridized or invaded
by brook trout, culverts could be replaced by intentional
barriers to protect these populations indefinitely. Good
barrier sites must be geomorphically stable and economically feasible, and the isolated stream network should
support a productive population and be large enough to
ensure long-term persistence. Larger networks and more
intact habitats would be higher priorities.
Figure 9—Opportunities and potential priorities for installing and removing barriers that isolate native
cutthroat trout in stream networks. The double parallel lines represent sites where large barriers might
be used to isolate relatively large stream networks that could protect socio-economic, evolutionary, and
some ecological values simultaneously (but see cautions about feasibility in the text). The four parallel
lines represent a similar site that might be a high priority because brook trout and rainbow X cutthroat
trout hybrids occur in nearby tributaries downstream. The black arrows point to existing culvert barriers
that are candidates either for conversion to intentional barriers to conserve remnant genetic diversity, or
for removal to expand connectivity and restore ecological and socio-economic values. All other barriers
above intentional barriers could be removed to expand the isolated network. The benefits of intentional
isolation would depend on the genetic integrity of the population above the existing barrier, the condition
of the habitat, and the ability to create a spatially diverse collection of isolated networks throughout the
basin. It would also depend on the feasibility of creating the barrier. Other existing barriers that are high
in the headwaters of tributaries are lower priorities for either permanent barriers or removal because
there is limited potential habitat above the site. Such small areas would either be vulnerable to local
extinction if isolated or contribute relatively little to the broader network if removed.
30
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
Because we are considering only evolutionary values
at this stage, representation, redundancy, and spreading of risk also become important criteria for selecting
sites for barriers. However, there has been no detailed
genetic inventory to guide management. Genetic similarity tends to decline with distance among salmonid
populations (Quinn 2005), and environmental conditions influencing local adaptation probably do also. If
the goal is to represent as much of the genetic diversity
and evolutionary legacy for westslope cutthroat trout
as possible, the logical objective would be to conserve
populations widely distributed throughout the system.
Broad representation would also minimize the threat
of simultaneous extinctions from large catastrophic
disturbances.
Barriers other than those selected above could be
removed to extend connectivity and conserve or restore
ecological and evolutionary values in other areas of the
basin. Barriers that currently isolate small fragments of
potential cutthroat trout habitat might be lower priorities
for removal than those that could expand the network,
regardless of the values in question (Figure 9).
Little Snake River
Background—Much different circumstances characterize the Little Snake River basin in southern
Wyoming and northern Colorado (Table 5, Figure 10).
The climate is typical of the central Rocky Mountains.
The first snowfall is in September or October, annual
Figure 10—Headwaters of the Little Snake River, showing the current estimated
distribution of populations of Colorado River cutthroat trout as bold stream segments
(after Hirsch and others 2006). In many cases stream segments shown as occupied are
fragmented by natural or artificial barriers, or the true distribution of native cutthroat trout
is incompletely known. Estimates of the lowermost extent of trout habitat in the Little
Snake River main stem in 1955 (Eiserman 1958) and in the Muddy Creek headwaters are
denoted by solid red bars.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
31
p­ recipitation peaks in March and April in the form of
snow, and there may be heavy rains associated with
monsoonal storms in July and August (Knight 1994).
Peak discharge is caused by snowmelt runoff in June,
with occasional localized stormflows in late summer.
Elevations range from 1900 m at the mouth of Muddy
Creek near Baggs, Wyoming, to over 3300 m in the
headwaters. Streams pass through three major vegetation types: sagebrush grasslands with riparian gallery
forests at the lowest elevations, extensive aspen stands
at mid-elevation sites, and coniferous forests higher in
the basin.
Land ownership and use are typical of many watersheds in the western U.S. Lower-elevation valleys are
privately owned and higher-elevation lands are in the
public domain. Historically this region served as a hub of
grazing activity. For example, in the early 1900s over 10
million sheep annually were pastured on summer range
or herded on stock driveways on the Wyoming side alone
(Thybony and others 1985). Livestock grazing continues
at much lower intensity, although sedimentation, riparian alteration, and irrigation diversions associated with
agriculture have altered fish habitat quality and tributary
access for decades (Eiserman 1958). Resource extraction
on public lands, such as hard rock mining and timber
harvest, is less than it was historically, but recreational
use of public lands has increased with population growth
in much of the mountain West, and publicly-built and
user-created roads and trails are extensive. An exception is a large portion of the uppermost North Fork
basin (Figure 10), which lies within the Huston Park
Wilderness Area. Finally, the entire North Fork basin
has undergone water development. Flows of all perennial
tributaries pass through water collection structures that
block upstream migrations of fish and fragment some
populations of cutthroat trout. In addition, portions of
the high spring flows (and probably juvenile fish) are
diverted into a pipeline transferring water east underneath the Continental Divide. Water development is also
projected or underway in adjacent Wyoming basins,
several of which harbor populations of cutthroat trout.
Native fishes are relatively few. Colorado River cutthroat trout once ranged throughout the basin, sympatric with mottled sculpin (Cottus bairdi) and mountain
whitefish in the colder upper reaches, and perhaps with
some members of the now-rare Colorado River fauna,
such as Colorado pikeminnow (Ptychocheilus lucius),
in the warmer downstream reaches of the main stem.
Eiserman (1958) considered the Little Snake River near
the mouth of Willow Creek and the town of Dixon as the
downstream extent of trout (Figure 10). Below this point
32
the river was regarded as too warm and silty to provide
trout habitat in summer, as were many of the lower
portions of Muddy Creek. Brook, brown, and rainbow
trout were first introduced in the 1930s (Eiserman 1958)
and currently dominate fish communities in coldwater
portions of the basin, whereas cutthroat trout are relegated almost solely to short stream segments protected
from upstream invasions by barriers. A waterfall near
its mouth protected the North Fork Little Snake River
population, one of the largest of this subspecies, from
early invasions. Later ones were thwarted by some of
the first barriers built (in the mid-1970s) specifically to
prevent upstream invasions of nonnative fishes. The main
stem has a mix of species, including nonnative channel
catfish (Ictalurus punctatus), but does not appear to
retain any native salmonids.
Presently, from 160 to 190 km of up to 38 streams
are believed to contain 15 to 20 distinct populations of
Colorado River cutthroat trout (Figure 10; Young and
others 1996, Hirsch and others 2006). These populations display different degrees of hybridization with
rainbow trout or Yellowstone cutthroat trout, but a few
populations appear to have retained their genetic integrity. Some populations in Colorado may have resulted
from early stocking of Colorado River cutthroat trout
originating from Trappers Lake, an adfluvial population in the headwaters of the White River basin in
northwestern Colorado. No lakes in the Little Snake
River basin were known to contain this subspecies.
Conservation values—Colorado River cutthroat
trout have been the subject of intensive management in
this basin (Speas and others 1994), primarily for their
evolutionary values. Their populations are rare enough
that each receives individual attention, and the focus of
management is to maintain all existing populations. We
infer from the extensive tributary network (for example,
Battle, Savery, and Slater creeks) that fluvial cutthroat
trout once existed. If so, their loss has reduced or eliminated ecological values such as demographic support of
resident cutthroat trout populations, transfer of nutrients
upstream, and providing food for terrestrial predators
like bears (Koel and others 2005). Socio-economic
values contribute little to the current management of
Colorado River cutthroat trout. Angling for cutthroat
trout is an incidental activity, primarily because their
populations are scattered, most streams are small and
difficult to access by automobile, and adult fish do not
reach large sizes because of the short growing season
at high elevations. Moreover, regulations prohibiting or
limiting harvest have been in place for decades.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
Vulnerability to invasion—Following decades of
stocking (Eiserman 1958, Wiltzius 1985), brook trout
and rainbow trout have become widely distributed in
most available habitats, and there is little evidence of
environmental or biotic resistance to their continued
spread should additional waters become accessible. For
example, within four years of an illegal introduction
just above the artificial barrier in the main-stem North
Fork Little Snake River, brook trout had spread to three
tributaries and ascended 8 km up the main stem despite
the robust population of Colorado River cutthroat trout
already present (M. Young, personal observation). In
addition, perennial streams lacking fish are exceptionally rare, and presumably contain geological or artificial
barriers, or habitat unsuitable for trout.
Persistence with isolation—Nearly all remaining
populations of Colorado River cutthroat trout persist
only above barriers (Figure 11), yet individuals in some
of these populations migrate hundreds or thousands of
meters to spawn (Young 1996) and move extensively
for foraging (Young and others 1997). Nevertheless
only four populations occupy more than 10 km of connected habitat and probably contain enough individuals
to afford security from the long-term loss of genetic
variation ­(Allendorf and others 1997, Young and others
2005). Most populations inhabit stream segments less
than 3 km (often due to natural or artificial barriers that
fragment habitat), likely consist of a few hundred fish at
most (Young and others 2005), and are at a heightened
risk of extirpation from an array of natural disturbances,
such as drought or post-wildfire debris torrents. Similar
populations of this subspecies, and salmonids elsewhere
in the inland West, have suffered this fate (M. Fowden,
Wyoming Game and Fish Department, personal communication; Brown and others 2001, Dunham and others
2003b).
Potential priorities and actions—Currently, most
populations of Colorado River cutthroat trout in this
basin inhabit short headwater stream segments and are
probably at an elevated risk of extirpation. The genetic
structure of these populations has not been studied, so
the most conservative approach would be to attempt to
Figure 11—In the Little Snake River watershed, Colorado River cutthroat trout persist only in headwater
streams, above barriers to upstream invasion by nonnative brook, brown, and rainbow trout (M. Young image).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
33
conserve all populations to maximize redundancy and
representation of whatever genetic diversity remains.
From this perspective, the smallest populations are at
the greatest risk. The probability of persistence of all
populations regardless of size could be enhanced by extending their distribution downstream to suitable habitats
currently occupied by nonnative species once the latter
have been removed. This approach could be particularly
valuable if it connects populations to additional streams
that decrease the risk of population loss from localized
catastrophic events. In addition, there may be opportunities to restore connectivity by providing fish passage
around water diversions, and to achieve redundancy by
introducing genetically pure populations into additional,
nearby waters that were previously occupied by cutthroat
trout. Such efforts are underway (B. Wengert, Wyoming
Game and Fish Department, and D. Brauch, Colorado
Division of Wildlife, personal communication), as are
attempts to identify waters where such actions could
be undertaken (Hirsch and others 2006). Nevertheless,
­attempts to establish populations in locations with
access to additional streams or to refugia that permit
natural recolonization has been, and will be, challenging.
Overcoming water management issues, resolving private
land-public land conflicts, and locating sites suitable for
barrier installation render this a long-term, site-specific
solution unless other methods for controlling nonnative
fishes become available. The feasibility of any of these
actions will likely guide the ultimate selection of sites
for further work.
Where suitable habitat exists but cutthroat trout
populations are absent, the primary conservation actions involve repeated chemical treatment or intensive
electrofishing to remove nonnative species and installing
artificial barriers to prevent reinvasions. This work has
been challenging. Occasionally, treatments have not
killed all nonnative fish present, forcing biologists to
repeat treatment of some streams or abandon reclamation in others. Some barriers were passable to upstream
migrating brook trout, or became so after damage by
high flows. In other cases brook trout were introduced
above these structures by anglers, apparently to retaliate for restricted harvest of cutthroat trout or the loss
of individually prized fisheries for nonnative trout. In
these latter cases, preservation of populations may rely
more on road closures, persistent public outreach, and a
regular law enforcement presence than on fish population management.
Finally, the relatively few indigenous populations of
Colorado River cutthroat trout remaining have served as
a source for fish stocked elsewhere in the basin, either
34
by direct transfers or indirectly as broodstock used in
hatcheries or off-site ponds. Although the use of lakes
as broodstock refugia has been successful elsewhere
(K. Rogers, Colorado Division of Wildlife, personal
communication), it has been repeatedly attempted with
limited success in this area. Use of hatchery or semi-wild
broodstocks argues for adopting genetic management
plans to ensure such stocks do not deviate from wild
populations (Mobrand and others 2005).
Section VII: Making Strategic _
Decisions_______________________
The framework we presented in Section V outlines
important questions to consider when managing the
invasion-isolation tradeoff. For biologists faced with
a particular basin and set of native salmonid populations that represent important conservation values, the
next step is to consider the available options and make
strategic decisions, before setting priorities for action.
A lesson from the two examples presented above is that
the best solutions for conserving native salmonids are
not the same in all regions. For example, in the Coeur
d’Alene River basin the main issues are whether nonnative salmonids will invade and displace or hybridize
with native cutthroat trout, and how much habitat will
be required to sustain important migratory life history
types if they are isolated. In contrast, in the Little Snake
River basin the primary issues are locating any remaining
pure populations, replicating them in headwater streams
above barriers after removing nonnative salmonids, and
improving the probability of persistence of these many
isolated populations.
Despite these fundamental differences among regions,
the two examples illustrate that fisheries managers do
have opportunities to make strategic decisions under
these different circumstances. For any given basin and
particular set of salmonid populations of conservation
value, these decisions will be influenced primarily
by two conditions: 1) the degree of isolation of the
populations of interest, and 2) the degree of invasion
threat (Figure 12). For example, if most native salmonid
populations in the basin are in small isolated patches, the
invasion of nonnative salmonids is either imminent or
relatively complete, and the invaders have strong negative effects (as in the Little Snake River basin; upper left
corner of Figure 12), then barriers will be required to
protect most populations from rapid extinction. Given
this, the strategic decisions available include: 1) the
size, habitat quality, and number of patches needed to
ensure persistence of a minimum number of the isolated
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
Figure 12—A conceptual diagram of the opportunities for strategic decisions when managing the
invasion – isolation tradeoff for native salmonid populations of conservation value. Examples of
strategic decisions to maximize conservation of remaining populations in several regions of the
decision space are shown.
p­ opulations through time; 2) the representation of as much
genetic, morphological, behavioral, and habitat diversity
as possible; 3) the spatial distribution and redundancy of
populations to avoid extinction of evolutionary lineages
by correlated catastrophic disturbances; and 4) working
out protocols for effective removal or control of nonnative trout, and subsequent translocations of native trout.
Despite these strong constraints on management options,
the goal of these decisions is to move the set of native
salmonid populations toward the lower right corner of
this decision space where the invasion threat is farther
away and patches are larger and more connected, even
though in many cases this goal is not fully achievable
at present.
At the opposite end of the spectrum, if a manager is
fortunate to have many large interconnected patches with
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
native salmonids, and the invasion threat is far away or the
effects of the invader are weak, then strategic decisions
will involve: 1) preventing invasions, 2) preventing fish
movement barriers or land uses that fragment habitat, 3)
monitoring the spread of any nonnative salmonids in the
basin, and 4) ensuring that natural processes continue
to provide the habitat heterogeneity that sustains the
native populations (Figure 12; lower right corner). For
example, a key strategy might be to minimize sources
of nonnative fishes that could foster unauthorized introductions (Rahel 2004) and educate the public about
their dangers to native biota (Cambray 2003a, 2003b).
Likewise, under this scenario managers have more opportunity to ensure representation and redundancy of
populations, and to experiment and monitor to learn and
thereby reduce future uncertainty.
35
Some managers will find themselves in other regions
of this decision space, where different strategies are
required. For example, a basin may have little invasion
threat but many culverts or diversions that fragment
headwater stream habitats (Figure 12; upper right
corner). In this situation, a strategic decision would be
to minimize habitat degradation (and improve habitat)
in these small patches to prevent further extinctions,
and increase connectivity by removing barriers where
feasible to maximize resilience of these populations. In
contrast, other managers may find that their basin has
large patches of habitat isolated by natural barriers, but
that nonnative trout were stocked above them and the
invasion is nearly complete (lower left corner). In this
case they may prioritize some of these basins for the
difficult task of removing nonnative trout, which may
require building temporary barriers to prevent upstream
invasion while successive treatments are used to extend
habitat downstream for native salmonids. This will require strong public relations and a long-term commitment
of agency funding and time. Overall, an understanding
of where their native salmonid populations lie in this
decision space will help managers develop strategic
decisions for other possible scenarios.
When barriers are used to prevent invasions, human
translocation of nonnative salmonids above them may
also be a problem, requiring further strategic decisions.
If native salmonids are restricted to many small patches
above barriers (Figure 12, upper left), managers could
construct several barriers at intervals (for example, 1
km) near the downstream end and monitor the buffer zone as insurance against reintroductions, but the
stream fragments are usually too short to justify other
measures. In contrast, in a large basin of intact and
relatively remote habitat where strong invaders are removed in stages (lower left), a strategic decision would
be to place the downstream barrier in an inaccessible
location to minimize human translocations. If temporary
barriers were constructed for the project, the materials
could be left on site to allow replacing barriers quickly
if translocations occur at a downstream barrier.
Finally, where constraints are strongest (Figure 12;
upper left corner), it is likely that managers will select a
mixed strategy and hedge their bets against uncertainty.
This may include conserving many small populations
that persist above barriers, translocating among replicates
when they go extinct (Lubow 1996, Hilderbrand 2002),
and working to protect or develop larger populations in
remote or protected basins (see Shepard and others 2005
for an example). In the worst case, if barriers on the large
populations are breached by human or natural agents
36
and invasions proceed quickly, fish from the smaller
replicate populations can be used to refound them.
Epilogue: What do We Need to
Know?
Conserving native salmonids at risk from invasions
in the western U.S. is a widespread, long-term, and
expensive problem, so investment in better information
to improve decisions is likely to pay off. In developing
this synthesis we found important uncertainties about
the key processes that drive the tradeoff between invasion and isolation. These often emerged as differences
in opinion among the authors and among reviewers,
based on their data and experience in different regions
and with different species and subspecies of salmonids.
From these discussions, we found three main questions
in need of research to inform better management:
• What are rules of thumb for stream length or watershed area needed to sustain different salmonid species
and subspecies in isolated populations above barriers? — It is clear that the size of the stream network
can strongly affect the resilience, persistence, and
viability of isolated salmonid populations. However,
these effects also vary strongly with environmental
conditions and the evolutionary history of species or
populations. Basic field data on species occurrence
and an inventory of natural and anthropogenic barriers can be used to develop models to estimate risk of
extinction with time (see Section III). Rapid progress
could be made by using existing data and collaborating with biologists already engaged in inventory and
monitoring programs throughout the region.
• What environmental, ecological, and evolutionary factors control where the main nonnative salmonids will
invade? — The extent and effects of nonnative invasions
appear to vary substantially across the interior West.
In many regions, it is not clear whether invasions are
continuing or stalled, and whether the process depends
on changing climate and even evolution of the species
involved. Management of invasions, and the effective use
of barriers and limited management resources, would
benefit from a better understanding of where and how
invasion and displacement is likely to progress.
• Does life history diversity of native salmonids buffer
invasion and allow coexistence? If so, how? — Conventional wisdom and life history theory suggest that
migratory forms could be more resilient and resistant
to invasion than resident life history types, due to differences in demographic processes and habitat use.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006
However, little is known about this topic so more specific
hypotheses and observations need to be developed,
followed by empirical analysis and simulation.
Although some initial research has been conducted on
each of these questions for specific species or environments (Dunham and Rieman 1999, Adams and others
2002, Harig and Fausch 2002), more is needed to allow
accurate predictions on which to base sound management.
Likewise, many other questions need to be answered
for a more complete understanding of the tradeoff (see
Dunham and others 2002a, Peterson and Fausch 2003b,
Dunham and others 2004) and the feasibility of different
management options. In addition, a careful monitoring
program followed by widespread reporting of the results
would rapidly advance our understanding of the need
for, and effects of, barriers used to manage populations
of native trout throughout the western U.S
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